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MOBILITY 0F SLUDGE-BORNE 3,3'
-DICHLOROBENZIDINE (DCB)

presented by

Duck Young Chung

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MOBILITY OF SLUDGE-BORNE

3,3'-DICHLOROBENZIDINE

BY

Duck Young Chung

A THESIS

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

MPSTER OF SCIENCE

Department of Crop and Soil Science

1987

ABSTRACT

MOBILITY 0F SLUDGE-BORNE 3,3'-DICHLOROBENZIDINE (DCB)

By

Duck Young Chung

Fate of the carcinogen 3,3-dichlorobenzidine (DCB) in sludge-
amended soil was evaluated. 14C-labelled 008 was used as a tracer to
examine mineralization and mobility of DCB present in sludge. DCB was
shown to be persistent in soil and sludge-amended soil with <2% of the
added [1401DCB evolved as laCOZ during a 182 day incubation. Mobility
of DCB was evaluated by using soil columns (5 x 45 cm) that were
irrigated with 2.54 cm water per day for 84 days. The total 14c
recovered in leachate water was from 0.16 to 0.063 percent of the
total 14C added to the top 15 cm of the soil column at time zero. The
strong binding of DCB to soil constituents rendered it highly
immobile despite being placed in a sandy soil which was irrigated
with large amounts of water. However, some finite, albeit small,
quantity was apparently leached through the soil profile at a
relatively constant rate. The immobilization of DCB presumably
results from the formation of humus bound DCB residues and from its

low water solubility

To my parents and family

ii

ACIQIOWLEL‘GZ'? TENTS

Sincere gratitude and appreciation is expressed to Dr. Stephen A.
Boyd for his encouragement, support and constructive criticism during
each phase of this project.

Gratitude is expressed to Dr. B. G. Ellis and Dr. A. J. M. Smucker
for their suggestions and support during this study.

I sincerely thank Mr. M; Mikesell for his help.

To Mi Suk, for encouragement and invaluable assistance I am deeply

grateful.

iii

Table of Contents

List of Tables ................................................ v
List of Figures ................................................ vii
Chapter I
Introduction ................................................... ,1
List of References ............................................. 24
Chapter II
Introductrion .................................................. 28
Experimental ................................................... 32
Materials ................................................. 32
Soil Incubation ........................................... 33
Soil Extraction .......................................... 34
Soil Columns .............................................. 35
14C Analysis .............................................. 36
Results and Discussion ......................................... 37
List of Reference ............................................ 50
Chapter III
Introduction ................................................... 52
Experimental ................................................... 60
Material.......... ........................................ 50
Soil column ............................................... 51
Planting and irrigation ................................... 51
Analysis ................................................ 52
Results and Discussion ......................................... 63

iv

List of Tables
Chapter I Page
1. Physical and chemical properties of 3,3'-dichlorobenzidine (DCB)4;
Chapter II
1. Mineralization of [IaC] DCB to 1[‘C02 during 182 days incubation in
sludge amended soil ........ : ...................................... 38
2. 14C distribution and recovery in soil columns and leachate ..... 48
Chapter III
1. Elemental analysis of treatment plant and sludge samples expressed
as range on a dry weight basis ....................... , .............. 55
2. Number of days required for a plant to develop from one stage to
the next ........................................................... 59
3. Total radioactivity in leachate water, soil and parts of soybean
plants ............................................................. 66
a. Dry weight of soybean plants grown in control and 1“C-DCB amended

80118 ............................................................. .69

List of References .............................. ' ............... 74

vi

List of Figures
Chapter I ~ Page

1. Proposed keto-enol tautomerism of dichlorobenzidine-based pigment,

with presumed ring structure.. ..................................... 3
2. Synthesis of DCB (3,3'7dichlorobenzidine DI ACID SULFATE ) ...... 5
3. Tetratization of benzidine ...................... . ................ 6
4. Synthesis of 0.1 pigment Yellow 12 .............................. 7
S. Imine formation reaction................ ........................ 15
6. Michael addition of aromatic amine to humus matrix .............. 16
7. Chemical structure of MC-DCA and HA-DCA ......................... 18
Chapter II

1. Chemical structure of 3,3'-Dichlorobenzidine .................... 29

2. Solvent extractable radioactivity from [IAC] DOB-amended soil and
sludge as a function of incubation time ............................. 39
3a. Distribution of radioactivity with depth in [IQC] DCB-amended soil
columns which were irrigated with 2.54 cm of D-H 0 per day for 84
days: 1(‘C -DCB plus soil ........................................ 44
3b. Distribution of Radioactivity with depth in [14C ]DCB-amended soil
columns which were irrigated with 2.54 cm of D-HZO per dayfor 84 days:
1(‘C-DCB plus sludge-amended soil ................................ 45
3c. Distribution of radioactivity with depth in [th]DCB~amended soil
columns which were irrigated with 2.54 cm of D-H20 per dayfor 84 days:
146 DCB amended sludge plus soil: Sludge rate was 10 g kg ....... 46
4. Cumulative 140 in leachate water from soil columns: (1)14C-DCB plus
soil, (2) 1(‘CuDCB plus sludge amendeds soil, and (3)140-DCB amended

sludge plus soil. Sludge rate was 10 g kg ....................... 47

vii

Chapter III Page

1. Distribution of radioactivity with depth in 1("C-DOB amended soil
columns which were irrigated with 30 mls of distilled water per day
for 18 days ........................................................ 64

2. Distribution of radioactivity in plant parts of soybean grown in

lZ‘C-DCB amended soil with sludge rates of 0 and 11.2 Mg/ha ......... 65

£14

3. Uptake o C-DCB by soybean plants grown in ll‘C-DCB contaminated

soil with sludge rates of 0 and 11.2 Mg/ha ............. ' ............ 71

14

4. Comparison of dry weight in parts of soybean plants grown in C-

DCB contaminated soil with sludge rates ofO and 11.2 Mg/ha ......... 72

S. 14C-DCB recovered as C02 respiration for 6 days after 11 days...73

Abbreviation:
Lzleaves szranch and stem Rzroot T:tota1

Sznon sludge-amended soil SS:sludge-amended soil C:control

CHAPTER I

PRODUCTION AND CHARACTERISTICS OF DCB AND SEWAGE SLUDGE, AND

BINDING OF AROMATIC AMINES TO SOIL AND HUMIC SUBSTANCES.

As a results of industrial discharges, sewage sludge generated
at the Muskegon Wastewater Management System (MWMS) has become
contaminated with the carcinogen 3,3'-dichlorobenzidine(DCB). The MWMS
uses wastewater to irrigate over 2,025 ha of land in Muskegon County.
Corn is the primary crop grown. Consideration is now being given to
initiate a program of land application of sewage sludge generated at
MWMS. Because MWMS sludge is contaminated with DCB, a potential hazard
of this sludge management alternative is for DCB to be leached from
sludge through the soil profile and into groundwater. The potential
for movement is enhanced because (1) DCB is known to be persistent in
agricultural soils, (2) the sandy soils at Muskegon have high
infiltration rates and the capability to percolate large amounts of
water and (3) these soils receive higher than normal quantities of
water during the year. For these reasons, a study was initiated to
evaluate the mobility of sludge-borne DCB in an experimental system
designed to mimic conditions at the MWMS.

3,3'-dichlorobenzidine (3,3'-dichloro,4,4'-diaminobipheny1),
hereinafter referred to as DCB, is widely used for producing 95

tetrazodyes listed in the Column Index (1971); however, of these, only

[\J

five pigments are currently produced in the United States. DCB is
suspected of being a human bladder carcinogen by the Occupational
Health and Safety Administration because of its similarity in
chemical structure to benzidine, a human bladder carcinogen (IARC,
1972a). It is of considerable commercial importance; DCB has been
used as a starting material for the synthesis of diarylide yellow and
orange azo pigments such as 5 chlorobenzidine-based pigments since
the DCB plant went on stream in 1938. Pigments yellow 12, 13 and 83
(DOB-based) are not carcinogenic in rats and mice. Pigments yellow 12
and 13 differ structurally from the benzidine-based dyes for which
there is positive evidence of carcinogenicity, which presumably
results from metabolic breakdown to release the parent benzidine
compound. In addition to a difference in solubility (pigments are
insoluble, whereas the dyes are water-soluble), DCB is a benzidine
derivative with chlorine substituents ortho to the amine. Figure 1
shows that a keto-enol tautomeric group adjacent to the azo linkage
in the pigments may shield this linkage from enzymatic cleavage
(Jones,l980).

The DOE pigments are used in printing inks, rubber, plastics,
paints and textiles. Total DCB production in the United States in
1977 was about 3.7 million pounds and approximately 1 million pounds
of DCB were purchased from an outside suppliers (Ferber, 1978).

Some physical and chemical properties of DCB ar shown Table l.
Chemically, DCB undergoes the reactions of the aromatic primary
amines. The most important is the formation of the tetrazonium

compound, which can be coupled with a variety of compounds for making

CH3 CH3
l
// .'. 3' \\
\ N :5 § N /
Leno”
CH: CH:
I I
= 0... Cl Cl H... 0 a
'§ ‘ N / \ N //
(~keto)

Figure 1. Proposed keto -enol tautomerism of
dichlorobenzidine -based pigment, with presumed
ring structure (EPA 560/ 3-78-003)

Table 1. Physical and Chemical Properties of
3,3;-Dichlorobenzidine ( DCB )

Synonym : 2,2'-Dichloro, [ l,l'-bipheny1 ]-4,4'-diamino

Chemical Formula :C12H10C12N2

Molecular Weight : 255.13 gram

Preparation from m-chloronitrobenzene

Needles from water or prisom alcohol

Melting Point : 165.5'C

Moderately soluble in alcohol

Readily soluble in ether

Insoluble in water

Soluble in ethanol and ethyl ether

pKa value : Approximately 4.4

Use : Manufacture of azo dyes

* Source : EPA 560/11-80-019 and Merck Index

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dyes and pigments. In these reactions the chlorine atoms of DCB are
passive and do not participate in the reaction. DCB is made by the
alkaline reduction of ortho-nitrochlorobenzene to
dichlorohydrazobenzene to DCB in the presence of hydrochloric or
sulfuric acid (Lurie, 1964). The product is isolated as the
dihydrochloride or as the acid sulfate. These reaction are shown in
Figure 2.

DCB reacts with sodium nitrite in the presence of
mineral acid ( HCl or H2804 ) to form the tetrazonium salt. The salt
is then coupled with various compounds, primarily
acetoacetarylides, to form the azo pigments (USEPA, 1979a). An
example of a coupling reaction is shown in Figure 3.

Production of dyes and pigments from dichlorobenzidine proceeds
via tetrazotization to form the tetrazonium salt, which is followed by
coupling of tetrazonium compound with a relative compound (e.g.,
aromatic hydroxy compounds or arylamines) to form a colored product
(1). Tetrazotization is accomplished by reacting benzidine or
dichlorobenzidine with nitrous acid (sodium nitrite in hydrochloric
acid) in a water solution at 0 ~5 C. This produces the coupling
agents, a tetrazonium hydrochloride (Fig.3).

The final step in the production of dye solution is the
coupling of the tetrazonium salt with phenol, aromatic amines, or
other reactive compounds(Fig.4). The coupled products have azo
groups that are linked to sp - hybridized carbon atoms; hence, they
are azo dyes. In the coupling reaction the first diazo group of the

tetrazonium salt of dichlorobenzidine couples readily to give the

diazomonoazo compound. The activity of the second diazonium group is
diminished so that formation of unsymmetrical diazo dyes also can be
accomplished (Lurie, 1964). By prOperchoice of coupling agents, the
resulting diazo compound can be diazotized or used itself as a
coupling component with another diazonium compound to give rise to
bisazo, triazo, tetrakis-azo, or poly azo dyes (Lurie, 1964). A wide
variety of dyes spanning the color spectrum is made possible by (l) a
large number of possible coupling agents, (2) the choice of
symmetrical or unsymmetrical coupling, and (3) the possibility of
forming monoazo or polyazo derivatives.

An important source of environmental release of dichloro-
benzidine is the wastewater from dye producer plants. Dichloro-
benzidine is very rapidly degraded in aqueous solution by the action
of natural or simulated sunlight (Sikka et al.,l978). Benzidine and
3~chlorobenzidine are intermediate in this process. In organic
solvents the dechlorination reaction is considerably slower (Sikka,
1978). D08 absorbs extensively to a variety of aquatic sediments and
becomes more tightly bound with time (Sikka,l978). Dichlorobenzidine
is more soluble in water and organic solvents than is DDT ( a well
known bioaccumulator ). DOB-based dyes and pigments discharged into
water from manufacturing facilities are believed to be chemically
reduced back to the parent compound if either hydrogen sulfide or
sulfur dioxide is present in the receiving waters. This was believed
to be the case downstream of a dye plant on the Sumida River in Japan
(Takemura et al., 1965). The DOB-based dyes and pigments discharged

into Sumida River from two manufacturing sites were accumulated in the

10
soil profiles, and gradually back to the parent compound such as
benzidine. Their degraduation rates were slower than the degradation
of other based dyes and pigments.

Hitz et al., (15) carried out laboratory studies on the
adsorption of dyes to activated sludge. After 30 minutes, average
adsorption of seven basic dyes ranged from 50 to 92 percent. No data
were found on the distribution of dichlorobenzidine in air. The rate
of DCB adsorption to all sediments studied by Gibson (1968) was
initially very rapid. The end point of sorption was generally
achieved within 24 hours (Gibson, 1968). Alkaline conditions may
change the properties of the sediments because treatment of sediments
with NaOH releases humic and fulvic acids from soil or sediments.

Metabolic conversion by aquatic microbial communities
represents one route for the ultimate degradation and elimination of
DCB residues. However, metabolites of aromatic amines may be more
toxic than the parent compound, thereby compounding the potential
hazard (IARC, 1972 a,b: IARC, 1973b: Jones, 1980). Appleton et al.,(2)
studied the degradation of DCB by aromatic microorganisms using
samples of wastewater and activated sludges. Their findings indicate
that DCB was not readily degraded by microorganisms. The progressive
decrease in DCB is probably related to the increase of organic
material. The action of sunlight on DCB leads to its degradation but
not necessary to its detoxification since benzidine, a relatively
photostable carcinogen, is one of the products (Gibson, 1968).

Boyd et al., (1984) studied the behavior of DCB in soil. They

observed that an initial sharp decrease in the amount of NaOH-

11

extractable DCB was accompanied by a sharp increase in NaOH-
extractable DCB of soil. This process indicated that movement of DCB
into humic fraction was essentially complete after 108 days since
NaOH-extractable DCB plus NaOH-extractable DCB remained constant in a
Brookston clay loam. Mineralization of DCB was very slow (ca. 2% per
year) in aerobic soil.

The formation of covalent linkages between DOB and soil humic
components appeared to be the primary fate of DCB. The behavior of DCB
in soil was consistent with that of structural analogues such as the
chloaniline which has been studied with respect to their binding
reaction and microbial transformation in soil (Bartha, 1971; Kearney
and Plimmer, 1972; Kaufman et al., 1973; Hsu and Bertha, l974a,b).

The fate of organic pollutant or xenobiotic chemicals in soil
has become a major health concern in recent times due to the potential
for surface and groundwater contamination and for crop uptake and
movement in the food chain. Certain xenobiotic chemicals pose serious
health hazards due to their toxic and carcinogenic properties. One
major environmental fate of hazardous organic chemicals is
immobilization at the soil surface by association with humic
substances. It is important to understand the nature of the
xenobiotic-humic substance complex, for only by understading this
association can we begin to put together a comprehensive soil
pollution indices that would relate relevant xenobiotic physiological
properties with bioavailability and hazard association data.

One important class of persistent xenobiotic compounds found in

soil are the alkyl- and halogen- substituted anilines. Aromatic

12
amines serves as the basic structural matrix for many commonly used
herbicides and pesticides and also used in the manufacture of plastic
dyes and pigments. The appearence of aromatic amines in the soil
environment may therefore be related to the intentional application of
herbicides or the unintentional releases often associated with
manufacturing process and waste disposal. Because the ring carbons are
not rapidly mineralized in soil(Bartha,197l), aromatic amines remain
available for binding reactions to soil humic substances. This
association with soil humus further stabilizes aromatic amines
against microbial degradation. Even chlorocatechols (produced from
degradation of 2,4-D) which are readily degradable in culture
solutions are stabilized in soil against microbial degradation by
attachment to soil humus (Statt et al.,l97l). Many aromatic compounds
are oxidized to catechols before the aromatic rings are cleaved
(Dagley, 1971).

It has been well documented that aromatic amines, when added
to soil, become strongly and irreversibly bound to soil humic
materials( Saxena and Eartha, 1983). This binding is believed to
result from the formation of covalent bonds between soil humus and the
amine group of anilines. Aromatic amines such as DCB form covalent
linkage to quinoidal sites of soil humus via nucleophilic addition of
the amine group.

When anilines are applied to soil, they are apparently bound
by both physical and chemical adsorption. It is also apparent that in
soil a dual absorption mechanism for aromatic amines is active. For

example, with dichloroaniline (DCA), a portion of the available DCA is

l3

formly (chemically) bound to the soil organic matter (Saxena,
l983).ther portionis loosely (physically) absorbed to both organic and
inorganic soil particles. The latter portion can be extracted by
organic solvents, the former portion can not. In whole soil the two
binding mechanisms compete with each other with the result that at a
given DCA concentration less DCA is bound firmly (chemically) to
whole soil than to a corresponding amount of isolated humic acid. With
time, a gradual shift occurs resulting in more chemically bound DCB.
Increasing DCA concentration resulted the increased amounts of bound
DCA, but the proportion of added DCA that become bound to soil or
humic acid gradually decreased with higher concentrations ( You and
Bartha, 1982 ).

Removal of the soil organic matter not only reduces overall
DCA-binding capacity of soil (Hsu and Bertha, 1976), but virtually
eliminates the type of binding that is resistant to solvent
extraction. The significance of bound residues as potential source of
future contamination is essentially unknown.

Several examples can be described which illustrate that for
aromatic amines in soil, binding processes are the prominent fate.
For example, an investigation by Chisaka and Kearney (9) demonstrated
that the herbicide 3',4'-dichloropropionanilide (propanil) is readily
transformed by soil microorganisms to 3,4-dichloroaniline (DCA) in
soil. They also observed that lacoz evolution from ring labelled
propanil ws less than 3 t after 25 days of incubatiuon in soil. Hsu
and Bertha (17) found that up to 90 t of the 3,4-dichloroaniiline

released during biodegradation of phenylamide herbicides becomes

l4

solvent-unextractable due to binding to soil organic matter. The
difference in bond stabilities for hydrolyzable versus non
hydrolyzable chloroaniline residue complexed with soil humic
substances suggested to Hsu and Bertha (19) that two distinctly
different covalent binding mechanisms were operative. Hsu and Bertha
also demonstrated that with time physically adsorbed and hydrolyzable
3,4-dichloroaniline residues slowly shifted to the non-
hydrolyzable form. Several investigators have studied the general
phenomenon of aromatic amine binding to soil humic substances. It is
generally thought that several different types of chemical or
biochemical reactions are responsible for binding of aromatic amines
to soil humic substances.

Hsu and Bertha (17,18) first proposed that imine formation to
explain rapid binding of primary aromatic amines to soil humic
substances (Fig.5). The reaction involving aromatic amines and the
carbonyl group of aldehyde or ketones to produce imines is reversible
reaction. The equilibrium can be shifted towards th two starting
vreactants if excess water is present. In soils, it is not clear
where the equilibrium lies under normal conditions (Parris, 1980).

There is, however, another type of binding reaction of aromatic
amines with soil humic substances that leads to the formation of
covalently bound residues. Cranwell and Haworth (10) and later Hsu
and Bertha (17,18) and Parris (23) proposed that primary or secondary
(aromatic) amines might undergo a nucleophillic substitution reaction
with a Michael acceptor. In this case the Michael acceptor would be a

quinone-like residue which is attached to the soil humic matrix

15

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R NR'

quuinone moiety in humote

Figure 6. 1 Michael addition of aromatic

amine to humus matrix, (Parris, 1980)

l7

(Fig.6). This is followed by tautomerization and oxidation yielding
aminoquinone (Reaction 2,3). The amino group may react further by
essentially the same addition -tautomerization-oxidation sequence to
form a variety of nitrogen heterocycles (Reaction 4). The nature of
the resulting heterocucle will depend upon the reactive sites (R) in
the humate juxtaposed to the aminoquinone. Progression along this
sequence of reactions is expected to make the amine moiety more
resistent to removal from the humate.

To date, three types of covalent aniline attachment to humus
and to model humic materials have been proposed: Formation of imines
,Micheel addition adducts (Hsu and Bertha l974a; Parris 1980), and
enzymatic oxidative cross-coupling reactions ( Bollag 1976 ). These
studies have demonstrated that the above type of model reactions
occured with incorporation of substituted anilines, and have
determined the chemical structures of the resulting products.

The comparison of hydrolysis and resistance of the 4-
methylcatechol-3,4-dichloaniline (MC-DCA) adduct with that of Humic
Acid-Dichloaniline (HA-DCA) complexes revaeled substantial differences
in behavior (You and Bertha, 1982). The MC-DCA adduct was completely
resistant to alkaline hydrolysis (Fig.7), but it was subject to
partial hydrolysis by acid. In addition, HA-DCA complexes were
partially hydrolyzed both by alkaline and by acid,the alkali being
somewhat more effective. In terms of biodegradation, the MC-DCA adduct
was found to be much more recalcitrant than HA-DCA complexes. If the
two attachment mechanisms were similar, one would expect the opposite

result, since structurally anallogous oligomers are generally more

CH3

 

 

 

Cl \
O
MC-DCA
O'H
(II-J \
-OH
HA-DCA

Figure 7. Chemical structure of MC -DCA
(4 -methylcatechol dichloroaniline) and HA -DCA
(Hydroquinone dichloroaniline).(You, LS. 1982)

l9

biodegradable than high molecular weight polymers (Albertsson and
Banhida, 1980). 14C02 release from HA-DCA complexes was stimulated by
the addition of aniline, butaniline failed to stimulate 14C02 release
from the MC-DCA adduct.This finding also supports the conclusion that
the bulk of DCA inhumic complexes is not likely to be bound in the
same manner asin the MC-DCA adduct ( You et al., 1982).

The imine formation and Michael addition mechanisms dicussed
thus far require, in principle, the presence of two reactants; an
aromatic amine and the carbonyl and quinone moieties of the
humic component. The next group of mechanism to be presented not only
requires the presence of these two reactants but also a biocatalyst
as well. In this process polyphenols are coupled together after
being enzymatically oxidized to the aryloxy radical: This enzyme may
also convert aromatic amines to radicals which then couple with soil
humic materials. Alternatively, the aromatic amine may react
chemically as described above with humic materials formed via
oxidative coupling reactions. Oxidetive coupling of polyphenols in
soils is though to be an important synthetic process leading to the
.formetion of humic substances. The possible source of polyphenols for
humus synthesis include lignin, microorganisms, uncombined phenols in
plants, glycosides, and tannins. Of these, the next three have
received attention: (1) Lignin, freed of its linkage with celluose
during decomposition of plant residues, is subjectes to oxidative
splitting with the formation of primary structural units (derivatives
of phenylpropane), (2) The side chains of the lignin-building units

are oxidized, demethylation occurs, and the resulting polyphenols are

20

converted to quinones by polyphenoloxidase enzymes. (3) Quinones
arising from the lignin (as well as from other sources) reacts with
N-containing compounds to form dark-colored polymers.

The results obtained by Bollag et al., (24) and Berry et al.,
(25) suggest that aromatic amine binding to soil humic components
results from a chemical reaction which may be only indirectly
controlled by the presence of oxidative coupling enzymes in soil. In
this, oxidative coupling enzymes act to create reactive quinoidal site
on the humic acid structure, and the amine reacts chemically via
nucleophillic addition reactions to form the bound residues.
Peroxidases and laccases are believed to be the responsible
catalysts in the coupling processes.

Aromatic amines may also be sorbed physically by soil.
Chemisorption may also occur if the amine moiety becomes protonated.
For the higher pH, sorption increase with increasing surface arae.
The sorption of benzidin by II'whole" soils and sedimnts has been known
to be controlled primarily by the concentration of the ionized
species. Sorption was highly correlated with pH, since pH controlled
the ratio of neutral to ionizd benzidine in the aqueous phase
( Zierath et al., 1980). The potential effects of benzidine and
related compounds can be greatly modified by sorption to soils and
sediments. However, the resistance of the aromatic amines residues to
solvent extraction and ion exchang combined with a sensitivity to acid
as well as alkaline hydrolysis strongly suggest a covalnt binding of
the nitrogn atom of these compounds to the carbon of carbonyl group or

to a quinoidal ring of the humic compounds.

21

As demonstrated with 3,4-dichloroaniline,' it is very
difficult to extract the anilines from soil by organic solvents or to
obtain their release by acid and alkaline hydrolysis (Bertha, 1971;
Chisaka and Kearney, 1970). It seems that anilines are bound to humus
or soil organic matter rather than to clay particles (Hsu and Bertha,
1974). It has usually been found that CO formation from halogenated
anilines occurs slowly (Bertha, 1971; Chisaka and Kearney, 1970), that
ring cleavages of these compounds by microorganisms is not a very
common reaction. The aeration by a continuous air stream and trapping
in an acidic solution reveal that most anilines are bound to a lesser
degree in autoclaved than in nonautoclaved soil ( Bollag et al.,
1983).

The absolute amount of DCA mineralized increased with
increasing DCA concentrations, although the percentage mineralized
decreased moderately ( You and Bertha, 1983). At low concentration,
the bulk of the aniline moiety becomes tightly bound to soil and
cannot be extracted by the salt, sediments or organic solvents
(Bertha, 1971; Chisaka and Kearney,l970). Physical absorption to
mineral soil particles competes with the chemical binding to humus,
but not only the chemically bound position resists exhaustive solvent
extraction.

In soil, DCB is immobilized rapidly by binding to humic
materials (Boyd et al., 1983). Boyd et al.,(7), observed that
essentially 90 t of the DCB added to soil was irreversibly bound to

soil humic fraction.

22

The progressive loss of solvent -extractable DCB, coupled with
increasing extractability (of the bound residue) by alkali, was
attributed to the formation of humus-bound DCB residues.
Mineralization of DCB in soil was very low (2 % in 6 months). The
formation of covalent bound DCB residues with soil humus is a natural
detoxication process in soils. The bgound residues are less mobile
and less bioavailable, and therefore less likely to contaminate water
or move into the food chain.

Berry and Boyd have studied ways to enhance the formation of
DCB-bound residues in soil as a decontamiunation method. A single
addition of ferulic and H202 to DCB-contaminaet soil significantly
decreasd extractable radioactivity and thus dramatically decreasd the
level of free DCB in soil. A previous study by Berry and Boyd (1984)
showed that phenolic compounds containing the three C-acrylic group
(e.g.,ferulic acid) were highly reactive with peroxidase enzymes, in
contrast to other naturally occuring polyphenols that did not contain
the acrylic group (e.g., vanillic acid). Addition of ferulic acid to a
reaction mixture containing peroxidase and chloaniline greatly
enhanced ( 2000K ) the rate of coupling of cxhloaniline: A similar
effect of ferulic acid on DCB binding in soil may occur.This procedure
may be usd in the future to enhance the covalent binding of DCB or
similar pollutants and thereby decrease the potential for movement
into ground water or the food chain.

The binding of xenobiotic chemicals to soil humic material is a
phenomenon which warrants an intense research effort both from the

standpoint of pure and applied science. It is important to understand

23

the physical nature of the binding process for only by undrstanding
the nature of the xenobiotic-humic substance interaction can we be in
aposition to access the contamination potential of soil and sludge-
born xenobiotics. It is also important to gain a fundamntal
understanding of these binding of xenobiotic compounds in soils.
enhancing the binding processes operative in soil which act to
covalently couple xenobiotic and humic substances may offer great

promise for detoxifying contaminatd soils.

REFERENCES

l. Albertson, A.C., and 2.6. Banhida. 1980. J. Appl. Polymer. Sci.
25:1656-1671.

2. Appleton, H.T., and Sikka, H.C. 1980. Accumulation,

elimination, and metabolism of dichlorobenzidine in the bluegill
sunfish. Envirn.Sci.Toxi. 14:50-54

3. Barthe,R,H.H. Blinke., and D. Prammer. 1968. Pesticide
transformations: Production of chloroazobenzenes from
chloroanilines.Sci. 16:582-583.

4. Barthe.R. 1971. Fate of herbicide-derived chloroanilines in soil. J.
Agr.Food. Chem. 19:380-381.

5. Bollag, J.M., R.D. Minard., and S.Y. Liu. 1983. Cross-linkage
between anilines and phenolic humus constituents. Envirn. Sci. Tech.
17:72-80.

6. Boyd S.A, Kao C.W, Suflita, J.M. 1983. Fate of 3,3'-DCB in
soilpersistence and binding. Environmental Toxicology and
Chemistry. Vol.3:201-208

7. Reaction rate of phenolic humus constituents and anilines during
cross coupling. Soil Biol. Bichem. 23:132-135.

8. Boyd, S.A.,C.W. Kao., and J.M. Sulfide. 1984. Fate of 3,3'-
dichlorobenzidine in soil:Persistence and binding. Envirn. Tox. Chem.
3:201-2089

9. Boyd S.A.,and D.F.Berry. 1984. Reaction rates of phenolic humus
constituents and aniline during cross coupling. Soil Bio1.Bichem.

23:132-135
24

25
10. Chisaka,H, end P.C. Kearney. 1970. Meyabolism of propanil in soils.

J. Agr.Food. Chem. 18:854-858.

11. Cranwell, D.A., and P.C. Kearney. 1970. Humic acid IV-The reaction
of amine acid esters with quinones. Tetrahedron. 27:1831-1837.

12. Dagley,S. 1971. Degradation of synthetic organic molecules in
biosphere. Proc. Conf. 1-15.

13. Ferber K.H. 1978. Benzidine and related biphenyldiamine.
Encyclopedia of Chemical Technology, 3rd ed., vol.3. New York:
Wiley/Interscience, pp. 727-777.

14. Gerarde H.W, Gerarde D.F. 1974. Industral experience with 3,3'-
dichlorobenzidine. Journal of Occupational Medicine. Vol. 16: 322-334
15. Gibson, D.T. 1968. Microbial degradation of aromatic compounds.
Science. 161:1093-1097.

16. Hsu, T.S., and R.Bartha. 1976. Hydrolyzable and non-hydrolyzable
3,4-dichloroaniline-humus complexes and their respective rates of
biodegerdation. J. Agr. Chem. 24:118-122.

17. Hsu, T.S., and Bertha, R. 1974. Interaction of pesticide-derived
chloroaniline residues with soil organic matter. Soil Sci. 116:444-452.
18. Hsu, T.S., and R. Bertha. 1974. Biodegradation of chloroaniline
humus complexs in soil and in culture solution. Soil.Sci 118:213-220.
19. Hitz H.R, Huber W, Reed R.H. 1978. The adsorption of dyes on
activated sludges. J. Society of Dyers and Chlorists 194:71-76

20. IARC. 1972b. International Association for Research on Cancer.
Cancer. Benzidine. IARC Monogr. 1:80-86.

21. Kaufman, D.D.,J.R.Plimmer, and U.J.Klingebiel. 1973. Microbial

oxidation of 4-chloroaniline. J. Agr.Chem. 21:127-132.

26
22. Kearney, p.c., and J.R.Plimmer. 1972. Metabolism of 3,4-

dichloroaniline in soils. J. Agr.Chem. 20:584-588.

23. Lurie A.P. 1964. Benzidine and related dieminobiphenyls. In:
Kirk RE, Othmer DF, eds. Encylopedia of Chemical Technology, 2nd
ed., vol 3. New York: Wiley/Interscience.

24. Miller, R.H. 1973. Soil microbial aspects of recycling sewage
sludges and waste effuluents on land. Soil SCi.114: 342-345.

25. Parris, G.E. 1980. Covalent binding of aromatic amines to humates.
1.reaction with carbonyls and quinones. Environ. Sci. Tech. 14:1099-
1106. 26. Saxena,A.,and Barthe.R. 1983. Microbial mineralization of
humic acidand 3,4-dichloroaniline complexes. Soil Biol. Bichem. 15:59-
62.

27. Sikka H.C, Appleton HT, Bennerjee S. 1978 Fate of 3,3'-
dichlorobenzidine in aquatic environments. Washington, DC:U.S.
Environmental Protention Agency. EPA 600/3-78-003.

28. Statt, D.E., J.P. Martin, D.D. Focht, and K. Haider. 1983.
Biodegraddation, stabilization in humus, and incorporation into soil
biomass of 2,4-D and chlorocatechol carbons. Soil Sci. Soc. Am. J.
47:66-70 29. Stevenson, F.J. 1982. Humus chemistry. Chap. 8:205-215.
30. Takemura H, Akiyama T, Nakajima C. 1965. A survey of the
pollution of the Sumida River, especially on the aromatic amines in the
water. Int. J. Air Water Pollutant (Great Britain) 9:665-670

31. You,I.S. and R.Bartha. 1982. Stimulation of 3,4-dichloroaniline
mineralization by aniline. Appl. Envirn. Microbial. 44:678-681.

32. You,I.S.,R.A.Jones., and R.Bartha. 1982. Evaluation of chemistry
defined model for the attachment of 3,4-DCA to humus. Bull. Envirn.

Contam. Toxicol. 29:476-482.

27
33. Zierath, J.J., Hassette., and W.R. Banwart. 1980. Sorption of

benzidine by sediments and soils. Soil. Sci. 129:277-281.

CHAPTER II

MOBILITY OF SLUDGE-BORNE 3,3'-DICHLOROBENZIDINE (DCB)

INTRODUCTION

3.3'-Dich1orobenzidine (3,3'-dichlo, 4,4'-diaminobipheny1),

hereafter referred to as DCB, is widely used as an intermediate in
the production of azo pigments. With current work practice,
effuluents containing this chemical are discharged directly into
receiving waters. Moreover, the discharge of DCB-pigment wastes
constitutes an additional source of DCB contamination in the
environment since free, unreacted DCB is reported to be present in
these pigments. As a result of industrial discharges like this,
sewage sludge generated at the Muskgon Wastewater Management System
(MWMS) has become contaminated with the carcinogen 3,3'-
dichlorobenzidine (Fig.1). The MWMS uses wastewater to irrigate over
2,025 ha of land in Muskegon County where corn is primary crop grown.

Consideration is now being given to initiate a program of land
application of sewage sludges generated at the MWMS. Because
agronomists have long reconized the value of sewage sludge in
improving soil conditions and supplying nutrients for plant growth.
Also, wastewaters that nomally are discharged into streams still
contain valuable nutrients and water resources that could be used for
agricultural production. However, applying sewage wastes to land as a
method of advanced wastewater treatment is not without its problem.

The MWMS sludge is contaminated with DCB, a potential hazard of this
28

 

3,3’-Dichlorobenzidine (DCB)

Figure 1. Chemical structure of 3, 3' dichlorobenzidine

30

sludge management alternative is for DCB to be leached from sludge
through the soil profile and into groundwaters.

Contamination of natural waters by a soil-borne xenobiotic
compound is a function of that compound's persistence and mobility in
the soil. For persistence compounds such as DCB, mobility in the soil
environment is specially important because the compound exists in
soil for long periods of time thereby enhancing the potential
movement and subsequent contamination of groundwaters. The potential
for movement is enhanced because (1) DCB is known to persistent in
agricultural soils, (2) the sandy soils at Muskegon County have high
infiltration rates and the capability to percolate large amounts of
water, and (3) these soils receive higher than normal quantities of
water during the year. For these reasons, a study was initiated to
evaluate the mobility of sludge-born DCB in an exprimental system
designed to mimic conditions at the MWMS.

DCB is produced in the United States and in several foreign
countries. Annual U.S. consumption of DCB reached 3.9 million pounds
in 1978 (Jones. 1980). The primary use for DCB is the production of
dyes and pigments. Although DCB can be used to produce 95 tetrazo
dyes, only five pigments are currently used in the United States:
Pigment Orange 13 and 14,and pigments. DCB is a demonstrated animal
carcinogen and has been shown to give a positive test in the Ames
Salmonella assay indicating that it possesses mutagenic properties
(Jones,l980). Furthermore, a positive correlation has been observed
between exposure levels of benzidine (in work place) and the
incidence of bladder cancer in humans (Jones,l980). the DCB based

pigment yellow 12 and 13 are not carcinogenic in rats and mice

31
(Jones,l980). The carcinogenic properties of other DCB-based
pigments are currently unknown.

The prominent fate of aromatic amines (such as DCB) in soil
generally appears to be a formation of stable covalent linkages with
soil humic substances (Boyd et al., 1984; Chisaka and Kearny, 1970;
Hsu and Bertha, 1974). Because aromatic amines are not readily
mineralized in soil they remain available for binding (Chisaka and
Kearny,l970; Hsu and Bertha,l974; Boyd et al.,1984). Generally, it
has been observed that after entry into the soil environment,
aromatic amines exhibit a sharp decrease in solvent-extractibility.
They showed that for DCB, loss of solvent-extractibility. Their
conclusion was that DCB had formed covalent linkages with the soil
humic fraction.

The formation of bound (non-extractable) residues rendered DCB
highly immobile (Boyd,1984), although some movement of DCB through
the soil profile was observed. In general, bound residue would be
expected to be less mobile and bioavailable than the unknown form.

Boyd et al., (1984) examined the fate of DCB in a Brookston clay
loam soil. Mineralization of DCB was very slow (ca. 2 % per year) in
aerobic soil. The behavior of DCB in soil was consistent with that of
structural analogues such as the chloroanilins which have been
studied with respect to their binding reactions and microbial
transformation in soil (Bertha,l97l; Kearny and Plimmer, 1972;
Keufeman et al.,l973; Hsu and Bertha, l974a; Parris,1980; and Saxna
and Bertha,l983).

In the present study, the mobility of sludge-born DCB in

sludge-amended soil was evaluated. Soil columns were constructed

32

using soil and DCB-contaminated sludge from the MWMS. Water was added
at a high rate of 2.54 cm per day for 84 days to simmulate the
irrigation schedule at Muskegon. Mineralization of sludge-borne DCB

in sludge amended soil was also examined.

EXPERIMENTAL

MATERIALS

A Rubicon sand (Entic Haplothod, sandy, mixed, frigid)
obtained at the Muskegon Wastewater Management System (MWMS)
Muskegon, Michigan was used. Soil was collected from the top 30cm of
the soil profile,air dried and passed through a 2 mm sieve. The
Michigan State University Soil Testing Laboratory carried out the
soil analysis with respect to particle size analyses, organic matter
content, pH and CEO. The Rubicon sand was found to contain the
following percent of sand was 93.5 percent; percent organic matter
was 2.54; pH was 7.7; and CEO was 6.6 me/lOO gram.

Sewage sludge contaminated with DCB was collected at the MWMS.
The sludge had originally been pumped from aeration cells and at the
time of sampling was in the form of a thick paste. The sludge was
spread out and allowed to air dry after which it was passed through a
2 mm sieve, and then stored- at 4 °C. Dichlorobenzidine content of
the sludge as 16 ug/g. Sludge was extracted with dichloromethane in a
Soxhlet extractor and quantitated for DCB using HPLC by the method of

Riggin and Howard (1979).

33

The 3,3'-dich1orobenzidine (3,3'-dichloro-4,4'diaminobiphenyl)
used in this study was purchased from Chem Service (West Chester, PA)
as the free base and was used without further purification.

DCB (ring-UL-la

C) was purchased from ICN Pharmaceuticals, Inc.,
a radiochemical purity of > 99% and a specific activity of
approximately 35 nCi/mmol. The [14C]DCB was dissolved in absolute
ethanol and stored in the dark at 0 °C. This gave a stock [14C]DCB
solution with a specific activity of 0.023 uCi/ uL and a DCB

concentration of 167 ug/mL.
SOIL INCUBATION

Mineralization of [14C]DCB was assayed in soil amended with
sludge at the rate of 0, 22.4, 112, and 224 Mg/ha (equivalent to 0,
10, 40 and 100 g sludge/kg soil). For each treatment, 50 g of air dry
soil and the appropriate amount of sludge were placed in 250 mL
Erlenmeyer flasks, and then moistened to approximately 60% field
moisture capacity with 4.5 mL distilled water. each flask then
received 40 uL of the [1401DCB stock solution (in ethanol) which was
added dropwise from a 50 ul hamilton syringe while the soil was being
stirred with spatula. In this fashion, 0.92 uCr of [1401DCB was added
to each flask. To trap any 1(‘COZ evolved, 1 mL of 0.1 N NaOH was
placed in a 1.4 mL plastic vial attached to a glass rod which extnded
downward from a rubber stopper used to close the flasks. Tropaolin O
indicator dye was added to the aqueous NaOH solution to indicat when
the NaOH had become saturated with 14C02. The flasks were opened

every 7 days (or sooner if necessary). at which time the 1[‘C02 traps

34

were collected and replaced with fresh aqueous NaOH. Flasks were

incubated in triplicate in the dark at room temperature for 182 days.
Soil Extraction

Change in the solvent extraction recovery of [14C] DCB over
time from soil and sludge were evaluated. In the recovery study, 500
g batches of soil or sludge were placed into large-mouth amber glass
jars. 0.5 milliliter of the [14C1DCB stock solution was added to 47.5
mL distilled water and mixed. Two 0.5 mL samples were withdrawn and
assyed for radioactivity. To each 500 g soil samplesl, 45 mL of the
aqueous [14C]DCB solution (60% field capacity) was added dropwise
using a Pasteur pipet while continuously stirring the soil. The jars
were closed and incubated for 28 days at room temperature. 35 g
samples were taken after 0.04, 0.17,1, 2, 5, 7, 10, 14, 21 and 28
days and stored frozen until analyzed. From each samples (35 g)
duplicate 10 g subsamples were placed in 125 mL Erlenmeyer flasks
along with 40 mL of a 60:40 (vol/vol) ethyl acetate/methanol solvent
mixture, closed with aluminum foil-coated rubber stoppers (No.5), and
allowed to shake for 22 hours on a rotary shaker at room temperature.
After sedimentation, 5 mL of solvent was sampled and centrifuged
(8000K g for 10 min.) from which 2 mL was used for liquid
scintillation counting. All extractions were performed in duplicate.
Figure 2 illustrates the amounts of solvent extractable radioactivity
and (expressed as a percentage of the added radioactivity) was
calculated on a soil or sludge air dry weight basis. Liquid-phase 14C

radioactivity was measured using a Beckman LS 8100 Liquid

35

Scintillation Counter. Samples were counted in Aquasol-Z (New England

Nuclear) LSC cocktail.
SOIL COLUMNS

Glass columns with teflon fittings and sintered glass bottoms
(Kontes) were used. The columns were 61 cm in length, had a 5.08 on
inner diameter, and were cut into 2.54 cm sctions from the bottom up.
The columns were taped together with plastic waterproof tape while
compressing the ends. The columns were then filled with 30.5 cm of
air dried soil and tapped repeatedly. The soil columns were then
immersed in a large cylinder filled with distilled water, allowed to
wet from the bottom up, and drained. This procdure was repeated twice
before adding the top 15.2 cm of soil or sludge-amended soil which
contained the [IAC]DCB. Two columns were prepared for each treatment.
The three treatments were: 1 mL of [1401DCB added directly to 410 g
soil with no sludge amendment, (2) 1 mL of [14C1DCB added to 410 g
soil amended with sludge at the rate of 11.2Mg/ha, (3) 1 mL of
[th]DCB added to 4.1 g sludge, incubated for 14 days to allow
equilibration on the added [14C1DCB, then 14C-sludge added to 410 g
of air dried soil.

For treatment 1 and 2, the soil or sludge amended soil was
throughly mixed and moistened with an aqueous solution (41 mL)
containing the [IACJDCB, incubated for 14 days in amber glass jars at
room temperature to allow equilibration of the added [IAC]DCB, and
then air dried. The air dried soil or sludge amended soil was then

added to the top of the soil column. Total soil depth was 45.7 cm.

36
each soil column received 35,000 nCr of [IACIDCB. Sludge was added at
a rate of 10 g/kg. Total weight of soil in the lower 30.5 cm was 820
g; the upper (treated) 15.2 cm soil zone weighed 410 g. At this point
(time zero), samples (4.1 g) of the [14C1DCB-soil or sludge amended

soils were taken and stored frozen for later 14

C analysis.

Distilled water (51.5 mL, equivalent to 2.54 cm) was added daily
for 84 days. Leachate water (50 mL) was collected in 500 mL
Erlenmeyer flasks (wrapped in foil) each week. A portion of this
water (100 mL) was extracted using a C18-Pak (Millipore) and eluted
with 2 mL of methnol into 10 mL scintillation cocktail. The pH of the
leachate was always >7. Trail extraction using leachate water spiked
with known quantities of [14C]DCB gave > 90% recoveries.

At the end of the 84 day irrigation period the columns were
disassembled into their 2.54 cm sections, air dried and weighed.
Distribution of the 14C with depth was determined by combusting
triplicate 1.0 g samples for each 2.54 cm section and trapping the
14C02 produced. The total amount of 14C added to each column was

determined by combusting samples of 14C-soil or sludge-amended

soil taken at time zero.
1“c ANALYSIS

Radioactivity was measured with a Beckman LSSlOO liquid
scintillation counter using external quench correction. 14002 samples
from the soil incubation and extraction studies were counted in
Aquasol-Z (New England Nuclear) LSC cocktail. Total soil 14C for

samples from the soil columns was determined using a OX200 Biological

37

1

Material Oxidizer (R.J. Harvy Instrument). The 4C02 was trapped and

14

counted in Carbo-Sorb (Research Product International Corp) LSC

cocktail.
RESULTS AND DISCUSSION

Mineralization of [14C]DCB was evaluated during a 182 day
incubation in soil amended with sludge at rates of 0, 10, 50 and 100
g/kg. The total amount of [140]DCB recovered as 1L‘COZ was less
than 2% in all cases demonstrating the persistence of DCB in soil
sludge-amended soil (Table 1). Mineralization of [17C]DCB in the
autoclaved controls was less by factor of at least 10X relative to
the nonautoclaved treatments. Thus, a biological component appeared
to be responsible for the slow release of 1“C02 from [1461DCB. The
effect of increasing sludge application rate was to increase the
total amount of DCB mineralized. However, the percent of DCB
converted to 002 decreased from 1.94% to 1.04% in going from a sludge
rate of 0 to 100 g/kg. These differences are quite small and the
important point is that DCB was highly resistant to mineralization in
both soil and sludge-amended soil

This result is consistent with a previous study showing the
persistence of DCB in a Brookston clay loam soil (Boyd et al., 1984).
The ability to recover added DCB from soil and sludge as a function
of incubation time was also evaluated (Fig. 2). Boyd et al.,(1984)
showed that the formation of soil bound (nonextrectable) residues
involving aromatic amines appears to occur rapidly. In this

experiment, a rapid and dramatic loss of solvent extractable 1(‘C

38

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Sludge

Radioactivity Extracted (96 of initial)

 

.—

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Incubation Time (days)

Figure 2. Solvent extractable radioactivity from [MC]
DCB-amended soil and sludge as a function of incubation
time.

40

from soil was observed during the first week of incubation. In this
study, a short term extraction experiment was conducted to examine
in greater detail the loss of solvent-extractable radioactivity
during the initial 4 weeks of incubation. The percent solvent-
extractable and the corresponding incubation times (in parentheses)
were 78 (1h), 73 (4h), 63 (1d), 52 (2d), 23 (5d), 18 (7d),15 (10d),
12 (14d), 11 (21d) and 9 (28d). After 7days incubation, approximately
18% of the added 14C could be extracted in soil after extraction with
ethyl acetate and methanol mixture. A more gradual decline in
extractable 14C occurred from day 7 to day 28 at which point
approximately 10% of the added 1‘0 was extractable. A similar loss
of solvent extractable 14C from sludge, which was observed, although
this loss was slower and less complete. The present solvent
extractable radioactivity and the corresponding incubation times (in
parentheses) were 74 (1h), 69 (4h), 64 (1d), 62 (2d), 51 (5d), 47
(7d), 35 (10d), 31 (14d), 23(21d) and 22 (28d). After 7days 47% of
the added 1“C was extracted and at the end of the incubation (28days)
22% was extracted. The greater recovery of added [1401DCB in sludge
was probably due to fewer binding sites for the formation of humus
bound DCB residues (Parris, 1980; Saxena and Bertha, 1983).

For aromatic amines in general, the time required for at least
50% of the residue to become bound is typically on the order of 1d to
1 wk (Hsu et al., Bollag and Voorman, 1981). Although some variance
exists in these initial rates, well over 50% of the residues
generally becomes nonextractable during the first several weeks of

incubation.

41

Recoveries after only 1 h incubation were quite good (74 and 78%
for soil and sludge) showing that the extraction procdure was
effective. Total 140 in soil and sludge was determined by combustion,
remained content over the 28 day incubation period (data not shown).
A similar loss of solvent extractable DCB, coupled with increasing
extractability by alkali, was observed previously for [laC]DCB added
to a Brookston clay loam soil (Boyd et al.,l984).

It has been well documented that aromatic amines, when added to
soil, become strongly and irreversibly bound to soil humic materials.
This binding results from the formation of covalent bonds between
soil humus and the amine group of anilines. For example, an
investigation by Chisaka and Kearny demonstrated that the herbicide
3',4'-dichloropropioanilid (propanil) is readily transformed to 3,4-
chloroaniline (DCA) in soil. there is, however, another type of
binding involving aromatic amines with soil humic substances that is
resistant to hydrolysis and not readily exchangable. Cranwell and
Haworth (1971) proposed that primary and secondary aromatic amines
might undergo a nucleophilic substitution with a Michael acceptor. In
this case, the Michael acceptor would be a quinone-like residue which
is attached to the soil humic matrix. The nucleophilic addition of a
primary or a scondary aromatic amine to the quinone moiety of the
soil humic substance matrix is a slow and reversible reaction which
is.followed by a rapid tautomeric step.

The formation of bound (nonextractable) DCB residues likely
results from the nucleophilic addition of the amine group to
quinoidal sites of soil humus (Hsu and Bertha, 1974a; Parris, 1980;

Saxena and Bertha,l983). By subsequent addition and oxidation

42
reactions the amine residue may become even more strongly bound to
the humate structure (Parris, 1980). The formation of a covalent bond
between DCB and soil organic matter strongly immobilizes DCB and, as
a result, the residues are no longer solvent extractable, which
accounts for the data in Figure 2.

Another immobilization mechanism operative in some soils is
protonation of DCB adsorption of this cation by clays. However, since
the first and second pKa's of DCB are less than 4 (Sikka et al.,
1978), DCB should exist primarily as the free base in the Rubicon
soil (pH 7.7).

Soil columns were used to directly assess leaching of [14C1DCB
in soil and sludge-amended soil. There are several features of the
column expriments which merit attention. The columns used had 5.08
on inner diameter and were 45.72 cm long. These dimensions were
selected to minimize the "edge effect" and to provide sufficient soil
column length to allow a natural soil moisture profile with depth.

The columns were irrigated with 2.54 cm of water per day which
was selected to exceed the amount of water irrigated in most, if not
all, wastewater irrigation system. This approach would, therefore,
identify the maximum DCB losses by leaching. The top 15.24 cm of the
columns were soil or sludge-amended soil which contained the
[14c10cs. Radiolablled [1401003 was uniformly distributed in the
upper 15.24 cm.

Distribution of 14c with depth after 84 days of irrigation is
illustrated in Figure 3. The vast majority of the 1[‘C-la‘belle
remained in the top 15.2 cm of the soil column in the

sludge-amended and nonsludge-amended soil. This attests to the

43
strong immobilization of DCB which showed little mobility despite
being placed in a sandy soil subjected to intensive irrigation with
2.54 cm of water per day. The distribution profile of 1(‘C from 15 to
45 on shows a finite and measurable quantity of radioactivity in each
2.4 on section of the bottom 30 cm of the soil columns (Table 2).
Radioactivity in the lower sections of the column tended to be higher
in non amended soil with sludge. Thus, these appears to be a very
"slow bleed" of DCB through the soil column. There does not appear to
be a band of 1“C radioactivity moving through the column.

Of the two sludge treatments (i.e. treatment 2 and 3), adding
the [IACJDCB to sludge, then incorporating the spiked sludge into
soil (treatment 3), resulted in less apparent DCB mobility. This may
in fact be more representative of actual field conditions then
treatment 2 where [140]DCB as added to sludge-amended soil

Radioactivity in leachate water collected from the soil columns
was very low (Figure 4 and Table 2) and all or part of this could be

140 in the bottom

due to lac impurity. However, when taken together,
30 cm of the soil column plus 146 in the leachate (Table 2) can not
be accounted for as radiochemical impurity which was < 1%. The total
amount of 1.[‘C in the leachate was from 0.16 to 0.063% of the total
14C added to the soil column (Table 2). Smaller amounts of total
radioactivity leached through the soil receiving sludge amendment
with treatment 3 showing the lowest 14C mobility

Figure 4 show the cumulative 14C-recovered in leachate water on
a weekly basis for 12 weeks. These data show that after about two

weeks the amount of 1"’C leached per week remained relatively constant

and showed no trend towards declining. These data suggest a “slow

44

 

 

 

 

 

 

 

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a .
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Figure 3a14Distribution of radioactivitywith
depth in [ CIDCB-amended soil columns
which were irrigated with 2.54cm of D-HZO
per day for 84 days: C-DCB plus soil.

45

 

 

 

 

 

 

 

 

 

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Tit-‘E'Qo
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dfiq'fiQQ'Q'tnNN
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Figure 3b. Distribution of radioactivity with
Depth in l' l4CIDCB-amended soil columns which
were irrigate with 2. 54cm of D-H 0 per day
for 84 days:1 C-DCB plus sludge- Inended soil.

Radioactivity (nCi/g)

100

99?]

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99.3]

90

997]
as 6]

80’

80.0]

70"

60-

50]-

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Figure 3c. Distribution of radioactivity
with depth in [MG IDCB-amended soil
columns which were irrigated with 2. 54cm
of H 0 per day for 84 days: l4C -DCB
amendedlsludge plus soil: Sludge rate was

10 gkg' .

Cummulative 14C in leachate (nCi)

47

 

 

l l

7 1421283542495663707784

Time (days)

Figure 4. Cumulative 1:5: in leachate water
fro soil columns: (I) C-DCB plus soil,
(2) MC-DCB plus sludge-amended soil, and
(3) C-DCB amended sludge plus soil.
Sludge rate was lOg kg '

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49

bleed" of 14C through the soil column. This movement,however small,

may continue for some extended period of time.

REFERENCES

l. Eartha, R. 1971. Fate of herbicide-derived chloroanilines in
soil. J. Agri. Food. Chem. 19:385-387.

2. Boyd. S.A., C.W. Kao, and J.M. Sulfita. 1984. Fate of 3,3'o
dichlorobenzidine in soil: Persistence and binding. Environ.
Toxicol. Chem. 3:201-208.

3. Cranwell, P.A. and R.D. Haworth. 1871. Tetrahedron. 27:1831-1837.
4. Hsu, T.S. and R. Bartha. 1974a Interactions of pesticide-derived
chloroaniline residues with soil organic matter. Soil Sci. 116:444-
452.

5. Hsu, T.S. and R. Bartha. l97hb. Biodegradation of chloroaniline
humus complexes in soil and in culture solution. Soil Sci. 118:213-
220.

6. IARC. 1972. International Association for Research on Cancer.
Benzidine. IARC Mongr. 1:80-86.

7. Jones, T.C. 1980. Benzidine, its congeners, and their derivative
dyes and pigments. Toxic Substances Control Act (TSCA) Chemical
Assessment Series. EPA-440/9-76-Ol8.

8. Kaufman, D.D., J.R.Plimmer, and U.I. Klingebiel. 1973.

Microbial oxidation of h-chloroaniline. J. Agri. Food. Chem. 21: 127-
132.

9. Kearney, P.C. and J.R. Plimmer. 1972. Metabolism of 3,4-
dichloroaniline in soils. J. Agri.Food. Chem. 20:584-585.

10. Parris, 0.3. 1980. Covalent binding of aromatic amines to
humates. 1.Reaction with carbonyls and quinones. Environ. Sci.
Technol. 14:1099-1106.

50

 

()I

ll. Ragin P.C. and A.B. Howard. 1979. Determination of benzidine,
dichlorobenzidine, and diphenylhydrazine in aqueous media by high
performance liquid chromatigraphy. Anal. Chem. 51: 210-214.

12. Sexana, A. and R.Bartha. 1983. Modeling the covalent attachment
of chloroaniline residues to quinoidal sites of soil humus. Bull.
Environ. Conta. Toxicol. 30:485-491.

13. Sikka, H.C., H.T. Appleton and S. Bannerjee. 1978. Fate of 3,3'-
dichlorobenzidine in aquatic environments. EPA 560/3-78-003. U.S.

Environmental Proctection Agency. Washington, D.C. ppz251-271

 

CHAPTER III

UPTAKE AND MOVEMENT OF 3,3'-DICHLOROBENZIDINE IN SOYBEAN PLANTS GROWN
IN SLUDGE-AMENDED SOIL AND NON SLEDGE-AMENDED SOIL CONTAMINATED WITH

[14c10c3

INTRODUCTION

3,3'dichlorobenzidine (3,3;-dichloro, 4,4'-diaminobiphenyl),
hereafter, referred to as DCB, is widely used as an intermediate in
the manufacture of azo pigments. It is of considerable commercial
importance; total DCB consumption in the United States in 1978 was
about 3.9 million pounds (Jones, 1980). Improper disposal of
industrial wastes containing the carcinogen DCB has resulted in the
contamination of soil, ground water near manufacturing sites. the
appearance of DCB as an environmental pollution, in addition to its
classification as carcinogen, demonstrates the importance of
understanding the fate of DCB in natural environment.

With current work practices, effuluents containing this
chemical are discharged directly into receiving water. Moreover, the
discharge of DCB-pigment wastes into receiving waters constitutes
additional source of DCB is reported to be present in theses
pigments. DCB is strongly carcinogenic to animals and regarded by
the Occupation Health and Safety Administration (OSHA) as being
carcinogenic to man. The discharge of DCB into aquatic environment is

of great concern to human health because of possible exposure to the

52

53

chemical through drinking water supplies. Also, if DCB and its
metabolites accumulate in the plants and food chains from
contaminated waters are to be used as human food, that could pose a
health hazard. The potential hazard of DCB may be compounded by its
biological or non-biological conversion to compounds of even greater
toxicity and/or persistence than the parent chemical. For instance,
it is believed that in the case of carcinogenic response. Further,
potential degradation products of DCB, such as benzidine, may
constitute an even greater carcinogenic hazards.

Currently nothing is known about the environmental fate of DCB
in a field planted crops. Therefore, this study was undertaken to
assess the role of some of the processes that may determine the fate
of DCB which is accumulated in soil andsludge-amended soil, leaching
rate by irrigation, influence of DCB on field grown soybean plants.
Particular attention was given to mobility of DCB and accumulation of
DCB in soybean plant parts. [IQCI-labelled DCB was used to trace
carbon through the various locations, including the final gaseous
form as respiration.

Sludge may be defined as a semi-liquid waste waste having a
suspended solid content of at least 2,500 ppm (i.e., 0.25 percent dry
solids) which flows, can be pumped, and exhibit delayed settling
characteristics in gravity settlers (Burd, 1968). Physically, sludge
generally has a thin, early odor likened by some to that of crude
oil. The chemical composition of sludge may vary considerably,
depending on the number and types of industries in the community, the
size and efficiency of the treatment plant, etc. Sludges are

separated and pressed in a variety of ways to convert them into a

54

stable condition for ultimate disposal. They may be classified
according to the stage of sewage treatment (e.g., primary, secondary,
and tertiary) or the specific process by which they were produced
(e.g., activated, chemically precipitated, digested, etc.). Municipal
sludges commonly certain an inorganic mineral fraction and a
combustible (i.e., organic matter) fraction. water usually makes up
90 percent or more of the weight.

Sludge contains many chemical elements in both mineral and
organic forms. Of the principal constituents, nitrogen (N),
phosphorus (P), and organic carbons are the most important in
promoting plant growth and good soil physical conditions. Potasium
(K) is present in only small quantities. Sampling and testing of
sludges to determine all the important constituents will be essential
before considering their use on agricultural land. Sludges vary
widely from community to community due to the source of wastewater
and proportion of industrial wastes they contain. Data in Table 1
illustrate how a large number of elements may vary among treatment.
plants handling essentially domestic and industrial flows, mostly
industrial flows. Values in Table 1 also illustrate the wide range in
elemental concentration in the sludges within the same wastewater
classification. This variability among plants which received similar
types of raw wastewater is not uncommon. In addition, elemental
concentration in the sludge from one particular treatment plant may
vary considerably with time. therefore, each sewage sludge treatment
should be carefully sampled and analyzed before considering its use
on land. Sludge contains appreciable amounts of N, P and small

quantities of K which are available for plant growth. So sludges can

U]
U!

have considerable value. Research and experiment have shown that
sludge can be beneficial both as a soil conditioner and as a low
analysis fertilizer. Corn, soybean, lawn and forage grasses, legumes,
and numerous ornamental annuls and perennials respond favorably. Some
vegetables, conifres, and other species do less well with sludge
fertilization because of lower nutrient requirements and greater
sensitivity to metals and salts. Sludge also provide organic matter,
which usually is beneficial as a soil conditioner. Organic matter
helps to hold plant nutrients in the soil which are then capable of
being released slowly during the growing season. Improved water-
holding characteristics, structure, and tilth generally result from
an increase in organic matter content, particularly for sandy soils
low in organic matter. On fine-textured soils, the organic additions
may result in improved infiltration capacities over time. while it
can never be a major contributor to the overall fertilization
requirements of crops grown in the field, it can provide important
amounts of nitrogen and phosphorus in localized area. As a soil
conditioner, sludges can promote desirable physical, chemical and
biological conditions in the soil for plant growth.

While plant nutrients and organic matter are useful to plants
and soils. Some of the metals and other constituents are potentially
harmful if not managed correctly. Table 1 shows the varieties and
wide range of elements found in sludge produced from treating
domestic and industrial wastewater. In addition to these elements,
other constituents may pose problems. Some concerns are; 1) Certain
metals can produce toxicity on less tolerant plant species; 2)

Additions of some potential toxic elements to soils may increase

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plant uptake of these elements and cause increased levels in the
human food chain; 3) A few metals, cadmium, for example, cam be
translated to the vegetative or grain portion of the plant and
thereby enter the food chain; 4) Organic contaminants like
polychlorinated biphenyl (PCB) or pesticides can be present in
sludges and might find a route into the food chain; 5) salts of many
kinds occur in sludges and can be deleterious to seed germination or
to growth of young plants.

Attention has been focused on utilizing sludge on crop land to
determine the rates which will optimize crop production. This method
uses sludges in a beneficial way and at rates will not result in (l)
excessive loss of NO -N to ground water; 2) high levels of P which
can eventually leach to ground water; 3) building up toxic levels of
metals in the soil; and 4) the bio-accumulation of elements in
plants which may be harmful to man or animals. To prevent these
concerns from becoming a problem, a number if factors must be
considered in determining proper application rates.

The fate and biological transport of [14C]DCB has been studied
in an effort to predict the extent of their transport in the food web
and their potential toxicity. Less is known of the effect of soil
concentration of DCB on availability to plants, especially at
background concentrations. plants vary in ability to accumulate
specific elements although physical and chemical processes in the
soil tend to be regulate the bioavailable fraction DCB in soil
solution.

The difference in uptake potential have been attributed to

physical variation in the development and intrusion of roots within

58

the soil and also to physiological differences such as not secretion
or the affinity of the membrane transport processes for DCB. Plant
age at sampling may affect total accumulation and changes in
observed concentration ratios that may occur as a result of
differences in rates of dry matter production at various stages of
plant development. Table 2 shows the number of days required for a
soybean plant to develop from one stage to the next. But less is
known about root growth corresponding to the number of days after
germination. In the great majority of species, seed germination
begins with radicle (embryomic root) rather than epicotyl (shoot)
protrusion through the seed coat (Berlyn, 1972). Root receive
photosynthataes and growth hormones from shoot and in return furnish
water and essential minerals. Root and shoot growth are thus
independent. During the first few days after wheat (Triticum aestivum
L.) seeds germinate the radicle grows faster than the plumule; this
relation causes the shoot-root (SzR) ratio by weight to be less than
1.0. Subsequently the plumule grows faster than the radicles and this
increases the shoot-root ratio (Aung, 1974). Soybean development can
be influenced by temperature , day length, variety, and other
factors.

Consequently, there can be considerable variation in the
number of days between stages. Temperature is the major factor
influencing vegetative development. Low temperature retards, and high
temperature enhances, seedling emergence and leaf development. The
effect of temperature becomes less important after the fifth-node
stage. A few node is produced on the main stem about every 3 days

after the fifth stage.

59

Table 2 . Number of days required for a plant to develop from one
stage to the next”

Stage Average number of days Range in number
of days *

Planting to VE 10 5-15

VE to VC 5 3-10

VC to V1 5 3-10

V1 to V2 5 3-10

V2 to V3 5 3-8

V3 to V4 5 3-8

V4 to V5 5 3-8

‘11 5°. 3C6 ..................... 3 ................. 7- -15 ............

Time interval between
all vegetative stages
after V5

*- Data are from studies conducted in the United States. The range
may differ from for cooler or more tropical climates.

60

The purpose of this study was to evaluate the uptake of soil-
deposited DCB using a plant species and soil type indigenous to an
area where DCB producing plants are located. Mobility of [140]DCB in
soil depth was evaluated after plants were planted for a short term

period.

MATERIAL AND METHODS

Pot experiment was conducted with soybean plant "Hardin" under
controlled environmental conditions (20 oC day and 10 0C night). The
pot which is polyvinyl chloride cylinders, having inside diameter of
7.6 and a wall thickness of 0.64 cm and containing 169 grams of sandy
loam in this experiment were used to establish the layered soil
containers was 2.5 cm. The containers were taped together with
plastic water proof tape while compressing the end. One end was
closed with cheese cloth to pass the irrigation water. The cores of
this study were filled a Rubicon sandy loam obtained at Muskegon,
Michigan. Soil was collected from the top 30 cm of soil profiles, air
dried and passed through a 2 mm sieve. Particle size analysis,
organic matter content, pH and CEO were determined by Michigan State
University Soil Testing Laboratory; percent sand, silt and clay was
82.4, 4.2, 13.4; percent organic matter was 2.45; pH was 7.7; and CEC
was 66 me/lOOg. Sewage sludge contaminated with DCB was collected at
MWMS. The sludge had originally been pumped from aeration cells and
at time of sampling was in the form of a thick paste. The sludge was

spread out and allowed to air dry after it was passed through a 2 mm

61

sieve, and then stored at 4 °C. Dichlorobenzidine content of the
sludge was 16 mg/ml.

3,3'-dichlorobenzidine, having a ring UL [14C], was
purchasedfrom ICN Pharmaceuticals, Inc., having a radiochemical
purity of > 99 t and a specific activity of approximately 35
nCi/mmole. The [laC]DCB was dissolved in absolute ethanol and stored
in the dark at 0 0C. This gave a stock [14C] solution with a specific
activity of 0.023 uCi [140] DCB/ul ethanol and a DCB concentration of

167/m1.
SOIL COLUMN

Polyvinyl chloride cylinders were used. The columns were then
filled with 5.08 cm of soil from the bottom and tapped repeatedly.
The soil columns were then immersed in a large cylinder filled with
distilled water, allowed to wet from the bottom up, and drained. The
procedure was repeated twice before adding the middle 2.54 cm of soil
or sludge-amended soil which contained the [14C]DCB, and top 2.54 cm
of soil. One column was prepared for sludge rate of 11.2 Mg/ha.
Sludge and soil were thoroughly mixed and moistened with and aqueous
solutions of 15 ul [1401033 and 19ml distilled water. These [IAC]DCB
amended soil were incubated for 6 days in amber jars at room
temperature to allow equilibration of the added [14C]DCB, then added
to the middle container for a soil depth of 2.54 cm. Total weight of
soil in the column was 480 grams and total weight of soil or sludge -
amended soil containing the [14C]DCB in the middle layer 2.54 cm was

160 grams. There were four replications of each treatment.

62

Sterilized soybean seed "Hardin" was used in this experiment.
Two soybean seeds were planted to each soil column directly at the
depth of 1.0 cm in the growth chamber. All soybean plants were
maintained in the growth chamber for 18 days. As irrigation
treatment, 30 mls of distilled water was added to the soil column for
18 days. Leachate water was collected in 500 mls of beaker for 18
days. The soybean plants were harvested after 18 days, and sampling
time corresponded to the second node stage. Sampled plants were
separated to leaves, branches, stems and roots. Sampled plants were
used for the following measurement; 1)Dry weight of each portion; 2)
amount of radioactivity in each portion. The entire portion of
leachate water was assayed for [IAC]DCB radioactivity.

After 18 days, the soil columns were disassembled into 2.54 cm
section, air dried and weighed. Distribution of the [IAC]DCB
radioactivity with depth was determined by combusting duplicate 1.0
gram samples from each 2.54 cm section and trapping the [14002]

produced.

ANALYSES

All laCOZ-radioactivity was measured using a Beckman LS 8100
liquid scintillation counter. Total 14602 for samples from soil
columns and soybean plants were determined using a OXZOO Biological
Material Oxidizer (R. J. Harvey Instrument). The 14C02 was trapped

and counted in Oxyflour-14C02 LSC Cocktail.

RESULT AND DISCUSSIONS

Soil columns were used to directly assess leaching of [14C]DCB
in soil and sludge-amended soil. There are several features of the
column experiments which merit attention. The columns used had a 2.54
cm and a 7.62 cm inner diameter. The soil columns were irrigated with
30 mls of distilled water per day which selected to provide enough
water for soybean growth from seedling to the second node stage. This
approach would, therefore, identify the maximum DCB losses by
leaching. The second layer of the soil column was amended with O and
11.2 Mg/ha of sludge contaminated with DCB. Radiolabelled [14C]DCB
was uniformly distributed at soil of the second layer.

Distribution of [14C]-radioactivity with depth after 18 days
of irrigation is illustrated in Fig.1. The vast majority of [14C1DCB
remained in the second layer cf the column in the sludge-amended and
non sludge-amended soil. This attests to the strong immobilization of
DCB which showed little mobility despite irrigation with 30 mls of
distilled water per day. There was, however, a finite and measurable
quantity of radiolabelled DCB in the first, third and fourth layer of
the soil column. Radioactivity in the lower section of the column
tended to be higher in soil that was not amended with sludge.
Radioactvity in leachate water collected from the soil column was
also measured (Table 3). The amount of [140] radioactivity in the
leachate was < 0.001 8. The higher amount of radioactivity leached
through the soil receiving no sludge amendment.

Uptake of [1401DCB by soybean roots and its translocation

throughout the soybean plants were evaluated. The general concept is

63

64

 

 

 

 

Soil Depth '
Figure 1. Distribution of radioactivity with depth

in 14C-DCB amended soil columns which were irrigated
with 30 mls of dostilled water per day for 18 days.

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that certain ions are adsorbed strongly by soil and hence are not
susceptible to rapid movement. The radioactivity transported by water
can be absorbed directly by the surface roots.

Plants vary in ability to accumulate specific radioactives
although physical and chemical processes in the soil solution. But
less is known of the effect of soil concentrations of DCB on
absorption to plants, especially at background concentration.

The relation between dry weight of roots and treatment is
shown Table 4. Dry weights of roots decreased in [14C]DCB amended
soil as follows;0.2937, 0.2934 and 0.4483.g/plant for sludge-amended
soil, non sludge-amended soil and control, respectively. At each
treatment, dry weight of roots decreased from the first layer to the
fourth layer. The decrease being greater in DCBoamended soil than in
control. The number of nodes were 3.8, 3.8 and 4.2 per plant in the
order of sludge-amended soil, non sludge-amended soil and control.
The length of roots ranged from 12 to 14.4 cm for all treatment. The
SzR ratio for sludge-amended soil, nonsludge-amended soil and control
was 1.53, 3.59,and 1.72, respectively. I concluded that non sludge-
amended soil contaminated with [14C1DCB increased S:R ratio of
soybean plants because of less immobilization of [140]DCB in non
sludge-amended soil.

Absorption of [14C1DCB in roots was 529.1 and 1028.26DPM in
sludgeoamended soil and non sludge-amended soil at rate of 0 and 11.2
Mg/ha, respectively. In each root layer, absorption of[IAC]DCB
concentration remained in soil columns, but not correspond to the dry
weight of root for each layer. In non sludge-amended soil, absorption

of [140]DCB of the second layer was not stronger than in sludge-

68

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Figure 4. Comparison of dry weight in parts of soybean
plants grown in C-DCB contaminated soil with sludge
rates of o and 11.2 Mg.ha'1.

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Figure 3. Uptake of l4C-DCB by soybean plants grown in
4C-DCB contaminated soil with sludge rates of 0 and
11.2 Mg.ha‘1'

72

amended soil. [14C]DCB concentration of roots for 18 days growth of
soybean plants was less in sludge-amended soil than in non sludge-
amended soil indicating that there was some stronger adsorption of
DCB to the sludge. Translocation of [14¢]DCB in plant parts, leaves
and branches, was 231.25 and 310.95 DPM for sludge-amended soil and
non sludge-amended soil. The higher uptake of [IACIDCB occurred in
non sludge-amended soil.

Total dry weight of soybean plants at each treatment were as
follows; 0.7407, 1.091 and 1.2323 g/plant for sludge-amended soil,
non sludge-amended soil and control,respectively. As we see in the
above results, yields were significantly and negatively affected by
the [IAC]DCB in the soil because dry weight of soybean plants grown
with [1&01008 and sludge at rate of 11.2 Mg/ha were smaller than
soybean plants with no sludge and no [14C]D08. And the concentration
of [140]DCB of plant parts for the fourth node stage were decreased
at the higher sludge application rate.

[th]radioactivity released as [laCOZ] respiration for 6 days
was observed. The range of [th]DCB released as respiration was 3.1
to 6.9DPM. this indicate that [140] radioactivity transported to the
plant tissues can also be released to the atmosphere even though the
amount of [140] radioactivity transported to the plant parts are very
small. In this experiment, the portion of [140]radioactivity released
as [14002] was <0.0001 s.

Total [140] radioactivity recovered from combustion was 93.49
and 82.55 t for in sludge-amended soil and non sludge-amended soil.
If sludge contaminated with DCB is utilized for land application, it

would be advisable to restrain from irrigation for some period time

73

to allow the immobilization of DCB to occur. Soil or sludge
contaminated with DCB affects growth of soybean plants. If there is
not a sludge-borne DCB, that will be less harmful. In most case,
sludge application to land will be used as fertilizer depending on
the contents of nutrients existed in sludge. DCB in the soil can move

upward or leach downward depending on irrigation.

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