Mutagenesis vol. 21 no. 5 pp. 295–304, 2006
Advance Access Publication 4 August 2006
doi:10.1093/mutage/gel037
The influence of environmental exposure to complex mixtures including PAHs and
lead on genotoxic effects in children living in Upper Silesia, Poland
Danuta Miel_zyńska*, Ewa Siwińska, Lucyna Kapka,
Krzysztof Szyfter1, Lisbeth E. Knudsen2 and
Domenico Franco Merlo3
Institute of Occupational Medicine and Environmental Health,
13 Kościelna Street, 41-200 Sosnowiec, Poland, 1Institute of Human
Genetics, Polish Academy of Sciences, 32 Strzeszynska Street,
60-479 Poznan, Poland, 2Institute of Public Health, University of
Copenhagen, Panum, Denmark and 3National Cancer Research Institute,
Largo R. Benzi 10, 16132 Genova, Italy
Environmental exposure is a complex mixture of hazardous compounds with different mechanisms of toxicity. In
case of concomitant exposure to carcinogenic substances—
such as polycyclic aromatic hydrocarbons (PAHs)—and
to heavy metals—such as lead (Pb)—the level of DNA
damage may be enhanced. Children are considered more
vulnerable than adults to chemical toxicants because they
take in more toxicants as a proportion of body mass and
because of inherent biological growth and developmental
factors. The objective of the study was to measure cytogenetic effects in Silesian children and to investigate their
relation with the environmental exposure to PAHs and Pb.
The examined population included 74 children 5–14-yearold who lived in two cities located in the most polluted
centre of the Silesia province. Individual exposure to lead
was assessed for each child by measuring lead in blood
(PbB), and to PAH by measuring 1-hydroxypyrene in
urine (1-OHP), urinary mutagenicity and DNA adducts
in circulating lymphocytes. Biomarkers of genetic effects
were assessed by measuring micronuclei (MN) and sister
chromatid exchanges (SCE) in children’s peripheral
lymphocytes. The mean levels of biomarkers of exposure
were as follows: PbB 7.69 mg/dl, DNA adducts 9.59 adducts
per 108 nt, 1-OHP 0.54 mmol/mol creatinine, and urinary
mutagenicity presented as the number of revertants per
mmol of creatinine: 485 for TA 98 and 1318 for YG1024.
Mean value of MN was 4.44 per 1000 binucleated cells and
SCE frequency ranged between 6.24 and 10.06 with a mean
value of 7.87. The results suggest the influence of exposure
to environmental agents on the induction of cytogenetic
effects in peripheral lymphocytes of children: namely Pb
on MN and PAHs on SCE. The sources of that exposure
may be outdoor and indoor. Emissions from coal-burning
stoves are important contributors to the total exposure to
PAHs and Pb in Silesian children.
Introduction
The Province of Silesia is located in the south of Poland that
occupies 2% of the total area of Poland and includes 11% of
the national population. The area is rich in natural resources,
mainly coal, zinc, lead and other ore mines, that have
stimulated the development of many industrial activities after
World War II which has caused an intensive deterioration
of the natural environment. The main sources of pollution are
mining and smelting activities and the presence of heavy
industry with coal-based power, steel and coke plants, as well
as heavy automobile traffic and use of coal for domestic
heating (1). Lead mining and processing have been operating
almost exclusively in this area of Poland. In the early 1990s
5 million tons of lead and zinc ores, corresponding to 4% of
the world total production, were extracted in the region. In the
late 1980s, annual lead fallout in the urban center of the
Province, known as Katowice Agglomeration, was 180 mg/m2
(2). In 1991 average annual concentration of benzo(a)pyrene of
airborne particulate matter from 24 sampling stations ranged
from 15 to 125 ng/m3, always exceeding the permissible
standard of 1 ng/m3 (average annual concentration) and
5 ng/m3 (average daily concentration) (3). The maximum
daily concentration was as high as 950 ng/m3 (4). Although in
the mid-1990s reduction of industrial activities reduced further
degradation of the environment, new hazards emerged,
including air pollutants from automobile exhaust and combustion of coal for domestic heating. Owing to the further
reduction of industrial activities the concentration of some
chemical air pollutants, including sulphur dioxide, nitrogen
dioxide and particulate matter, has further decreased in recent
years. By 2004 annual lead fallout had gradually decreased
with only 13 measurement points out of 707 exceeding the
permissible level of 100 mg/m2. Conversely annual concentrations of benzo(a)pyrene were still exceeding the standard value
of 1 ng/m3 in all sampling stations of the District SanitaryEpidemiological Station in Katowice with values as high as
18.6 ng/m3 (5).
It has been hypothesized that children are more susceptible
than adults to environmental toxins (6–8). Such susceptibility
can be attributed to the fact that they have greater exposures to
environmental toxicants than adults in relation to their body
weight, they undergo rapid growth and development and their
developmental processes are easily disrupted (9). Children
have also a longer life span to develop chronic diseases that
may be triggered by early exposures. Chronic diseases (e.g.
cancer) that are caused by environmental toxicants are thought
to arise through a series of stages that require years or even
decades to evolve from their initiation to clinical manifestation.
Carcinogenic and toxic exposures sustained early in life
appear more likely to lead to disease than similar exposures
encountered later (9).
Environmental exposure to lead in childhood was and still
remains a relevant environmental health problem in Silesia
Province (1,10). Given its genotoxic potential (11–21),
research focused on the identification and specification of
DNA damage in children exposed to lead is of particular
importance. Children play in leaded dust and dirt and have
*To whom correspondence should be addressed. Tel: þ48 32 266 08 85 (ext. 271); Fax: þ48 32 266 11 24; Email: [email protected]
The Author 2006. Published by Oxford University Press on behalf of the UK Environmental Mutagen Society.
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295
D.Miel_zyńska et al.
more hand-to-mouth activity than adults and they have higher
gastrointestinal absorption once they have ingested lead (1).
Blood lead levels as low as 10 mg/dl can lower subsequent
scores on IQ tests in children whereas the same exposure would
not be expected to be of much consequence in adults (22).
The evidence of lead carcinogenicity remains unclear. Lead
and lead compounds have been reviewed three times by the
International Agency for Research on Cancer (IARC) (23–25).
Currently, lead and inorganic lead compounds are classified
by IARC in Group 2B, possibly carcinogenic to humans, on
the basis of sufficient evidence for carcinogenicity in
experimental animals. Exposure of cells in culture has shown
that lead can induce chromosomal aberration, micronuclei
(MN) and sister-chromatid exchange (SCE) (13,26) but the
biological mechanism is still unknown (18). The evidence for
chromosomal toxicity of lead has been reported inconsistently
in exposed humans (10,11,14,17,19,26,27). A few studies on
DNA damage were carried out in children environmentally
exposed to lead (27–31) and their results remain inconclusive.
Polycyclic aromatic hydrocarbons (PAHs) are the most
prominent among the genotoxic and carcinogenic agents
present in polluted urban air. They have been the focus
of attention of most air pollution-related biomarker field
studies including biomarkers of exposure like 1-hydroxypyrene
(1-OHP) (32) or DNA adducts (33) and biomarkers of
genotoxicity (34). Urinary mutagenicity reflecting global
exposure to a mixture of known and unknown mutagens/
carcinogens has also been used as an unspecific biomarker of
exposure to PAHs (35,36). Both epidemiological and experimental evidence indicate that PAHs are carcinogenic to
animals and possibly to human beings (25). The mechanism
of genotoxicity of PAHs is well known and many studies on the
genotoxic effect of airborne PAHs in humans have been carried
out. They refer both to occupational (37–40) and environmental exposure including children and adolescents (30,41,42).
Environmental exposure is a complex mixture of hazardous
compounds with different mechanisms of toxicity. Hazardous
pollutants contained in ambient air are the main source of
environmental pollution in Silesia region. PAHs and their
nitro-derivatives are responsible for the high mutagenic
activity of ambient air particles in Silesia (43,44). Lead is
suspected to affect the efficacy of the repair mechanisms in
case of DNA damage caused by other agents, including PAHs
(18). Concomitant exposure to these chemical compounds is
still a very important problem in this region of Poland.
The objective of the study was to detect cytogenetic effects
in Silesian children and to investigate their relation with the
environmental exposure to PAHs and lead.
Materials and methods
Study design
The examined population included 74 children (47 boys and 27 girls): 5–14year-old who lived in Katowice and Sosnowiec, two cities located in the most
polluted centre of the Silesia province. The study was approved by the Ethics
Committee of the Silesian Medical Academy in Katowice. According to the
Polish law parents of all children participating in this study gave their written
consent.
All parents completed a self-administered questionnaire form including
items concerning children’s demographic variables (age, gender) health status
(past and present diseases including chronic diseases), current and past intake
of medicines, and dietary habits, including consumption of fruits and
vegetables grown in Silesia Province. For each child included in the study
parent’s education and smoking habit (including number of cigarettes smoked
daily) were collected. Information concerning X-ray exposure (diagnostic,
296
dentistry) and the presence of cancer within the child’s family and among
parent’s first-order relatives was also collected. Other questionnaire items
included the presence of industrial activities, the intensity of traffic at the
dwelling site and the type of heating system (electric, oil, gas, coal stoves).
Individual exposure was assessed for each child by measuring lead in blood
(PbB), 1-OHP in urine, PAH–DNA adducts in circulating lymphocytes
and urinary mutagenicity. Biomarkers of genetic effects were assessed by
measuring MN and SCE in children’s peripheral lymphocytes.
Blood and urine sampling
Blood sampling took place in the outpatient clinic of environmental health at
the Institute of Occupational Medicine and Environmental Health in Sosnowiec
during the cold season ranging between November (1998) and March (1999)
when heating is required. Blood samples were collected into sodium–heparin
tubes—Vacuette for setting cell cultures and Vacutainer for PbB determination.
Twenty-four hour urine samples were collected into sterilized plastic
containers.
PbB determination
Lead in whole blood was determined by electrothermal atomic-absorption
spectrophotometry (EAAS) according to Stoeppler et al. (45). Vortex-mixed
blood (200 ml) was added to 800 ml of 5% HNO3 in a pre-cleaned 2.2 ml
Eppendorf tube. The mixture was vortexed and left for 24 h in refrigerator for
better deproteinization. After centrifuging at 10 000 r.p.m. for 15 min the
supernatant was transferred to the Perkin–Elmer polystyrene autosampler cups.
The solution (20 ml) was automatically injected into the pyro-coated graphite
tube with L’vov platform. The Pb in the sample was vaporized at the optimized
sequential dry-atomize furnace program (Transversely Heated Graphite
Atomizer, THGA) developed in the laboratory. The atomic absorption signal
of lead was measured in the absorbance-peak area mode using Zeeman effect
for background correction (Perkin–Elmer 4100ZL). The amount of lead in the
blood samples was calculated by reference to matrix-matched calibration plots.
The detection limit of PbB was 6 mg/l.
Determination of 1-OHP in urine
The determination of 1-OHP in urine was carried out using the high-pressure
liquid chromatography (HPLC) method developed by Jongeneelen and Anzion
(46). 1-OHP in the 20 ml urine samples was enzymatically deconjugated and
then transferred to primed C18 Octadecyl cartridges (J.T.Baker), washed with
10 ml of water, and eluted with 9 ml of Methanol. The components of the eluate
were separated by means of HPLC on the HP 1090 apparatus (Hewlett Packard)
with the ODS C18, 200 · 2.1 mm column whereas 1-OHP was quantitatively
determined using the fluorescence detector HP 1046 (Hewlett Packard) with
229 nm excitation and 400 nm emission wavelengths.
The detection limit was 0.14 nmol of 1-OHP/l of urine. 1-OHP concentrations were expressed in mmol/mol creatinine to account for differences in
urine dilution.
Determination of creatinine in urine
Urinary creatinine concentrations were analyzed using a standard colorimetric
method following the picric acid reaction and absorption at 520 nm (47).
Urinary mutagenicity
The analysis included the preparation of urine extracts according to Yamasaki
and Ames (48). The organic substances present in urine were condensed using
adsorption and desorption methods on columns filled with organic resin
XAD-2. The extracts were examined by the aid of Ames incorporation plate test
(49). Before the examination the acetone extracts of urine were dissolved in
DMSO (dimethyl sulfoxide).
Three doses, 20, 40 and 80 ml of DMSO solution (representing 6, 12 and
24 ml of urine), were examined twice with Salmonella typhimurium strains:
TA98 and its derivative with elevated levels of O-acetyltransferase enzymes:
YG1024, both with metabolic activation (þS9 mixture).
The significance of linear dose–response relationship was examined by the
least square method with the assumption of P-level 0.05. Mutagenic effect
was expressed as a number of induced revertants per mmol of creatinine (50).
Determination of DNA adducts in white blood cells
Peripheral blood (10 ml) was collected into heparinized tubes and stored in
20 C till total WBC DNA extraction but not longer than 1 month.
DNA extraction
DNA isolation was carried out using proteinase K digestion, phenol–
chloroform extraction and ethanol precipitation as described in standard
protocols. The concentration of DNA was determined spectrophotometrically
by measuring the UV absorbance at 260 nm and the purity was ascertained by
the ratio at 260/280 nm.
Genotoxicity in environmentally exposed children
Table I. Selected characteristics of the Silesian children population by gender
Silesian
children
Girls
Boys
Total
N
27
47
74
Age in years
[mean (SD)]
8.9 (2.3)
8.0 (1.7)
8.4 (2.0)
Exposure to
ETS [N (%)]
15 (60.0)
27 (60.9)
42 (60.6)
Presence of coal
stove in home [N (%)]
9 (34.6)
15 (34.1)
24 (34.3)
Parents’ educationa [N (%)]
Primary
Secondary or higher
14 (58.3)
22 (53.7)
36 (55.4)
10 (41.7)
19 (46.3)
29 (44.6)
a
Not available for nine children.
Aromatic DNA adducts analysis
DNA adducts were assessed using 32P-postlabelling method mainly described
by Randerath and Randerath (51). DNA (5 mg) was digested with micrococcal
nuclease, phosphodiesterase from calf spleen and nuclease P1 from Penicillum
citrinum (the last one is used to enrich the mixture in modified nucleotides).
Then the digested was labeled with [g-32P]ATP catalyzed by T4 polynucleotide
kinase. Redundant [g-32P]ATP were eliminated by apyrase. The labeled material was then separated by thin layer chromatography on polythyleneimine–
cellulose plates and estimated by autoradiography. The diagonally radioactive
zone was excized and quantified by scintillation counting. The parts of the
plates without spots were used as background.
Determination of cytogenetic endpoints in lymphocytes
Cultures. Venous blood was taken from each subject using heparinized
vacutainer tubes. Lymphocyte cultures were set up by adding 0.5 ml of
heparinized blood to 4.5 ml of medium (RPMI 1640; Gibco) supplemented with
20% heat-inactivated fetal bovine serum (FBS; Gibco), antibiotics (penicillin
and streptomycin) and L-glutamine. Lymphocytes were stimulated by 1%
phytohaemagglutinin (Gibco).
MN assay. The cultures were incubated at 37 C for 72 and 44 h after
the initiation of cultures, cytochalasin-B (Sigma) at a concentration of 6 mg/ml
was added to arrest cytokinesis. MN slides were stained with 10% Giemsa
in phosphate buffer. A total of 1000 binucleated cells with well preserved
cytoplasm was examined for each subject on coded slides and the total
number of MN and the frequency of binucleated cells with MN (BNMN) were
scored (52).
SCE assay. The cultures were incubated for 72 h at 37 C with 0.25 ml of
5-bromo-20 -deoxyuridine (BrdU; Sigma). Colcemid (Gibco) was added 2 h
before harvesting. The cells were collected by centrifugation, resuspended in a
pre-warmed hypotonic solution (0.075 M KCl) for 20 min and fixed in acetic
acid/methanol (1:3, v/v). Chromosome preparations and stained slides were
prepared following the Anthosina and Poriadkowa procedure (53). SCE were
scored in 50 metaphases and presented as a number of SCE per cell. Data from
<50 metaphases were discarded (54).
Statistical analysis
Questionnaire and analytical data were stored in a database and statistically
analyzed using STATISTICA for Windows, Version 9.9, 1997. Biomarkers
data were log transformed to meet normality assumptions if required by the
test statistics (comparison of mean values). Test results were considered
statistically significant for the P-value <0.05.
The levels of SCE were normally distributed (Shapiro–Wilks test). The
distribution of urinary 1-OHP concentration and the levels of DNA adducts and
MN were skewed to the right. Therefore, they were transformed as log (Xþ1) to
make their distribution normal (1-OHP and DNA adducts) or stabilize the
variance (MN). The arithmetic mean and the standard deviation as well as the
median and the range were used to describe the frequency distribution of
biomarkers of exposure (PbB, 1-OHP and DNA adducts) and effects [MN and
SCE in peripheral blood lymphocytes (PBL)]. Univariate statistics (Student’s
t or Mann–Whitney tests) were used to compare subgroups dichotomized
according to individual covariates including gender (boys and girls), age
(<9 and 9 years old), exposure to ETS (yes ¼ 1, no ¼ 0), exposure to indoor
emission (coal stoves inside flat ¼ 1, other heating system ¼ 0), and parents’
education (defined as the highest level of education achieved by the mother
or the father: primary education ¼ 1, secondary or higher education ¼ 0), PbB
levels (<10, 10 mg/dl). Correlations between biomarkers of exposure and
effects were calculated by Spearman’s rank test.
Multiple linear regression analysis (stepwise procedure) was used to assess
the influence of PbB, white blood cells DNA adducts and urinary 1-OHP—used
as continuous covariates—on MN and SCE levels separately. Gender, age,
exposure to ETS and/or indoor emissions due to coal stoves and parents’
education were included in the regression model.
Table II. Levels of biomarkers of exposure and genetic damage in Silesian
children
Biomarker/assay (unit)
Biomarkers of exposure
PbB (mg/dl)
DNA adducts (per 108 nt)
1-OHP (mmol/mol creatinine)
Urine mutagenicity
(revertants/mmol creatinine)
TA98þS9
YG 1024þS9
Biomarkers of eect
MN (MN/1000)
SCEs (SCE/cell)
No.
Mean (SD)
Median
Range
74
72
66
7.69 (4.29)
9.59 (7.97)
0.54 (0.31)
6.77
7.17
0.46
2.7–23.0
0.64–44.92
0.04–1.95
65
64
485 (415)
1318 (909)
398
1021
95–2934
45–4987
71
72
4.44 (3.76)
7.87 (0.87)
3
7.86
0–17
6.24–10.06
Results
A total of 74 children, 5–14-year-old, participated in the study
between November 1998 and March 1999 (Table I). Examined
children included 47 boys (63.5%) and 27 girls (36.5%). Mean
age of children was 8.4 years. Of the children 60.6% were
exposed to ETS by co-habitation with smokers (mother,
father or both), and 34.3% were exposed to indoor pollution
generated by the system of heating based on coal-burning
stoves. Primary education was achieved by the parents of
36 children and secondary or higher level by the parents of 29.
For 9 children parent’s education was not available (Table I).
Table II shows the average levels of biomarkers of exposure
measured in children’s blood (PbB and DNA adducts), and
in urine (1-OHP and urine mutagenicity). The mean PbB concentration was 7.69 mg/dl, with levels as high as 23.0 mg/dl.
DNA adducts were detected in all children ranging between
0.64 and 44.92 adducts per 108 nt and with mean level of 9.59
adducts per 108 nt. 1-OHP mean value was 0.54 mmol/mol
creatinine with highest values close to 2 mmol/mol creatinine.
The mean level of revertants per mmol of creatinine corrected
for urine dilution was 485 for TA98þS9 (range 95–2934) and
1318 (range 45–4987) for YG1024þS9.
MN and SCE measured in PBL are reported in Table II.
MN mean value was 4.44 per 1000 binucleated cells, with
individual values reaching 17 MN cells per 1000 binucleated
cells. SCE frequency ranged between 6.24 and 10.06 with a
mean value of 7.87.
Correlation between biomarkers of exposure and effects
was examined by Spearman’s correlation. Positive significant
correlation was found between the level of SCE and PAH–
DNA adducts (r ¼ 0.250, P < 0.05) (Figure 1) and between
PbB and MN levels (r ¼ 0.347, P < 0.05) (Figure 2). The mean
level of MN in children with PbB concentration exceeding the
threshold PbB level (10 mg/dl) was significantly higher than
297
D.Miel_zyńska et al.
Fig. 1. Spearman correlation between the number of MN in PBL and concentrations of PbB.
Fig. 2. Spearman correlation between number of SCE and DNA adducts in PBL.
in children with PbB level equal or below the threshold value
(6.4 versus 3.9, respectively, P < 0.05; Fig. 3). SCE mean
levels were not found to be associated with the PbB threshold
(data not shown). Significant correlation was also found
between biomarkers of exposure: 1-OHP and DNA adducts
(r ¼ 0.471, P < 0.05) and 1-OHP versus urinary mutagenicity
by YG1024 (r ¼ 0.291, P < 0.05).
298
We examined the influence of gender, age, indoor exposure
to ETS, emissions from indoor coal burning stoves, as well as
parents’ education on the level of biomarkers measured in
children (Table III). Blood lead concentration was significantly
affected by age, ETS, emission from coal stoves and parents’
education. Increased PbB levels were found in children who
were at least 9 years old, exposed to ETS and to indoor
Genotoxicity in environmentally exposed children
Fig. 3. Box and whiskers plot showing the number of MN cells/103 binucleated cells in children with PbB level 10 mg/dl and >10 mg/dl.
Table III. Influence of selected covariates on biomarkers of exposure and effects
Covariates (n)
Gender
Boys (47)
Girls (27)
Age
<9 (52)
9 (20)
ETSa
None (26)
Yes (40)
Coal stove
No (46)
Yes (24)
Parents’ education
Low (36)
Median (29)
Pb
(mg/dl)
Mean (SD)***
1-OHP
(mmol/mol creat.)
Mean (SD)***
TA98þS9
(rev./mmol creat.)
Mean (SD)***
YG1024þS9
(rev./mmol creat.)
Mean (SD)***
DNA adducts
(per 108 nt)
Mean (SD)***
MN
(per 1000 cells)
Mean (SD)**
SCE
(per cell)
Mean (SD)#
8.43 (4.85)
6.40 (2.67)
0.51 (0.29)
0.58 (0.35)
458 (237)
529 (600)
1286 (835)
1369 (1030)
8.54 (8.72)*
11.47 (6.17)
4.91 (3.44)
3.67 (4.16)
7.63 (0.83)*
8.28 (0.78)
6.43 (3.90)*
9.25 (4.20)
0.57 (0.34)
0.48 (0.20)
521 (453)
396 (317)
1262 (816)
1410 (1166)
9.40 (6.90)
10.39 (9.50)
3.46 (3.09)*
5.60 (4.22)
8.05 (0.92)
7.69 (0.70)
6.18 (3.19)*
8.69 (4.80)
0.51 (0.40)
0.56 (0.25)
552 (633)
470 (238)
1455 (1144)
1292 (787)
8.27 (6.53)
9.40 (6.83)
4.64 (3.19)
4.36 (4.11)
7.95 (0.92)
7.78 (0.86)
6.83 (3.95)*
9.39 (4.70)
0.43 (0.21)*
0.74 (0.36)
387 (241)*
674 (589)
1064 (750)*
1804 (1002)
8.15 (7.07)
10.68 (6.30)
4.02 (3.95)
5.30 (3.31)
7.77 (0.88)
7.98 (0.87)
8.89 (4.83)*
5.61 (2.82)
0.63 (0.35)*
0.43 (0.20)
582 (505)*
368 (266)
1529 (932)*
1045 (888)
9.57 (6.83)
8.18 (7.09)
4.77 (3.70)
4.33 (4.12)
7.85 (0.97)
7.85 (0.87)
a
None or one or both parents; *P < 0.050; **Mann–Whitney U-test; ***Student’s t-test (performed on log transformed data); #Student’s t-test
(performed on original data).
emission from coal operating stoves. Increased PbB concentrations were detected in children whose parents had a primary
education compared to those with secondary or higher
education. The influence of gender was not statistically
significant although PbB level in boys tended to be higher
than in girls (Table III). Indoor coal burning stoves and parents’
education affected the level of 1-OHP and urinary mutagenicity detected by both bacterial strains: TA98þS9 and
YG1024þS9, resulting in increased concentration of these
biomarkers in children exposed to emissions from coal stove
and in children whose parents were low educated. The levels
of DNA adducts and SCE were clearly and significantly
associated with gender, with higher levels of these biomarkers
in girls than in boys. Increased levels of DNA adducts were
detected among children exposed to indoor emissions from
coal stoves although the difference was not statistically
significant. ETS exposure was not associated with increased
DNA levels. The highest level of DNA adducts was detected
among girls exposed to indoor coal stove emissions (13.17 per
108 nt SD ¼ 4.87, data not shown). The level of MN was
significantly higher in boys than in girls and in older (9 years
old) than in younger (<9 years) children. MN mean level was
299
D.Miel_zyńska et al.
Table IV. Determinants of MN and SCE levels in children’s PBL
(multiple linear regression analysis)
Dependent variable
Independent variables
MN/1000 cells
PbB (mg/dl)
Gender
ETS exposure
DNA adducts (per 108 nt)
Gender
ETS exposure
SCE/cell
b
SE (b)
P-value
0.355
0.076
0.254
0.234
0.305
0.114
0.118
0.115
0.119
0.128
0.131
0.116
0.007
0.536
0.046
0.063
0.016
0.330
b, standardized regression coefficient; SE, standard error; adjusted for age,
coal stove’s indoor emissions and parents’ education.
increased in children exposed to indoor emissions from coal
burning stoves although the difference was not statistically
significant.
The relationship between the level of PbB, 1-OHP and DNA
adducts on the number of MN and SCE in PBL was also tested
using multiple linear regression analysis, taking into account
all confounding factors (Table IV). As far as MN were considered the only clearly significant association was found for
PbB level (b ¼ 0.335, P ¼ 0.007). Exposure to ETS was
inversely associated with the number of MN (b ¼ 0.254,
P ¼ 0.046). SCE were statistically affected by gender (higher
in girls) (b ¼ 0.305, P ¼ 0.016). Despite the lack of statistical
significance DNA adducts were found to be associated with
SCE (b ¼ 0.234, P ¼ 0.063).
Discussion
Children are considered highly vulnerable to chemical
toxicants because of potentially higher exposure caused by
higher intake as a proportion of body mass and higher
absorption compared with adults as well as inherent biological
growth and developmental factors (8). It is therefore important
to investigate whether exposure to common and ubiquitous
environmental pollutants, such as Pb and PAHs may result in
genotoxic effects in exposed children.
Initiatives aimed at the prevention of lead poisoning in
Polish children began in the early 1980s. They were focused on
children living close to industrial sources of lead. In 1993 the
Institute of Occupational Medicine and Environmental Health
(Sosnowiec, Poland), began a childhood lead poisoning
prevention program oriented to children living in the urban
center of Silesia, beyond the direct impact of industrial sources
of lead (1). It was based on the determination of the most
important biomarker of the exposure to lead i.e. lead in whole
blood concentration. Kasznia-Kocot et al. (10) determined the
level of PbB in 2–4-year-old children attending nursery school
in three districts of the city. The geometric mean was 13.1,
13.2 and 11.9 mg/dl depending on the district, with the highest
observed value of 30 mg/dl. Concentrations >20 mg/dl were
found in 19.5, 11.8 and 15% of the children. In the period
of 1993–1998, 11 877 Polish children aged 24–84 months
living in six cities of Silesia were examined for PbB
concentration. The geometric mean of PbB was 6.3 mg/dl
(range 0.6–48 mg/dl), and >13% of them had PbB level
10 mg/dl. Poor housing, lack of recreational trips outside
Silesia, and the time spent outdoors were associated with
elevated PbB (1). Similar or even higher concentrations of
PbB were detected in our research as well as in 222 children
aged 11–15 years from Portoscuso, Sardinia, Italy (55) and in
300
131 schoolchildren 6–8 years old from Jakarta (56). The boys
of Portoscuso had a mean blood lead level of 11.30 mg/dl
and girls of 7.39 mg/dl. Mean PbB concentration of Jakarta
children was 8.3 ± 2.8 and 6.9 ± 3.5 mg/dl, depending on the
region and 26.7% of children had PbB level 10 mg/dl. Much
lower PbB levels were found in 1994 among 1230 boys
and 1211 girls, aged between 3 and 19 years (median age ¼
10 years) from the south of Sweden. The geometric mean was
2.7 mg/dl (range 1.2–12.2 mg/dl) in boys and 2.3 mg/dl (range
1.2–9.7 mg/dl) in girls (57). In our study geometric mean of
PbB concentration was 6.7 mg/dl with the values ranging
between 2.7 and 23 mg/dl. PbB level exceeded current
threshold of concern of 10 mg/dl in 21.6% of Silesian children.
PbB levels were higher in children aged 9 years, in children
exposed to indoor emissions from coal-operated stoves and
ETS, and whose parents were with low education. There was
no significant effect of gender on the PbB level, however,
PbB concentration tended to be higher in boys than in girls.
High concentrations of PAHs in ambient air are also
substantial environmental problem in Silesia province. Biomarkers of PAH exposure that are available for use in
molecular epidemiology studies include adducts with DNA
urinary metabolites like urinary 1-OHP and urinary mutagenicity. The concentration of 1-OHP in children was
measured in three Silesian regions in late 1990s. Median
concentration of 1-OHP in 7–8-year-old children not exposed
to ETS (n ¼ 126) was 0.32 mmol/mol creatinine and children
exposed to ETS and living in the town of Bytom, an old
urbanized area with an out-dated industry (n ¼ 108) was
0.49 mmol/mol creatinine (58). Median 1-OHP level in 30,
8-year-old children from the town of Sosnowiec, Silesia
province, sampled in July on 6 consecutive days ranged from
0.28 to 0.59 mmol/mol creatinine (59) with the highest value
reaching 3.62 mmol/mol creatinine. In our study median
concentration of 1-OHP was 0.46 mmol/mol creatinine with
highest values close to 2 mmol/mol creatinine and arithmetic
mean was 0.54 ± 0.31 mmol/mol creatinine. Concentrations of
1-OHP in Silesian children are high as compared with two
reports on urinary 1-OHP levels in children from Japan and
The Netherlands. In the study involving 644 Dutch children
the mean 1-OHP creatinine adjusted concentration was
0.34 mmol/mol creatinine (60). Japanese authors did not
correct the values of 1-OHP for creatinine concentration.
Mean levels of 1-OHP in urine samples collected in October
from 10-year-old Japanese children attending elementary
schools in three urban areas ranged between 0.46 and
0.73 nmol/l (61), whereas mean value found in Polish children
was 2.72 nmol/l. The relatively high concentration of 1-OHP
in Polish children reflects high environmental exposure to
PAHs in Silesia province. The results of the study confirmed
our previous findings that indoor pollution resulting from
coal heating and/or cooking strongly affects the level of urinary
1-OHP (58).
Besides one paper (62), there are no published research
papers on urinary mutagenicity in children. Moreover, different
methods of preparation of urine extracts made inter-research
comparisons very difficult. In order to make the comparison
possible we recalculated our data from the number of revertants
per mmol creatinine into mutagenic rate (MR/24 ml). The
mutagenic rate (MR) is a ratio of a number of induced
revertants to a number of revertants in the control sample.
Median mutagenic rate of urinary mutagenicity in Silesian
children determined with TA98þS9 is at the same level as in
Genotoxicity in environmentally exposed children
not occupationally exposed men from Zabrze (62) and
D˛a browa Górnicza (63): 1.46, 1.4 and 1.32, respectively, but
higher than in non-smoking men: 1.2 (55) and 0.66 (63).
We did not observe the differences in urinary mutagenicity
(expressed as a number of revertants in 24 ml) detected by
TA98þS9 and YG1024þS9 between population of children
and non-smoking women from Silesia (environmental exposure) and Białystok (control group) (64). Median values for the
strain TA98 were: 27 versus 24.1 versus 20.9 and for YG1024:
130 versus 101.6 versus 84.9, respectively. Urinary mutagenicity in no-smoking men living in two Silesian towns (35)
was slightly lower than in Silesian children. Mutagenic rate
(MR/24 ml) for TA98 was 0.9 versus 1.42 and for YG1024:
2.8 versus 3.06. The only research on urinary mutagenicity in
children concerned boys and girls living in central Poland, close
to the landfill. They had the same MR/24 ml of urine tested by
TA98þS9 (1.4 versus 1.42) (65) as our children but significantly
lower in case of the strain YG1024: 1.46 versus 3.07
(D. Miel_zyńska, E. Siwińska, unpublished data). As in the case
of PbB and urinary 1-OHP concentration, the level of urinary
mutagenicity was influenced by coal stoves emissions as well as
by parents’ education. The influence of the level of education may
be connected with worse living conditions of low educated people
who live in old buildings and use coal burning stoves placed
in rooms and kitchens for heating and cooking purposes.
DNA adducts are commonly employed biomarkers in
molecular epidemiology. The great majority of studies focused
on DNA adducts is related to tobacco smoking and occupational exposures. There are very few studies investigating
DNA adducts in children including newborns. DNA adducts
in nasal respiratory epithelium were significantly increased in
86 children, 6–13-year-old, exposed to urban air pollutants in
the Metropolitan Mexico City, compared with 12 unexposed
children aged 11 years (66). A significant increase of PAH–
DNA adducts was also detected in cord blood from newborns
of unemployed mothers who lived in polluted areas of the city
of Krakow, Poland (n ¼ 17) compared with six newborns of
mothers from low polluted areas (8.14 versus 1.7 adducts/
108 nt) (67,68). Aromatic DNA adduct level was studied in
various subpopulations living in Silesia region (64). For
example, a level of bulky DNA adducts equal to 5.0/108 nt
was established in leukocytes of people living in vicinity of
coke oven plant (69). Our findings showed a 30% increased
level of DNA adducts in children exposed to indoor emissions
from coal stoves. Conversely, ETS was not associated with
higher DNA adducts levels suggesting a major role of the
exposure to coal stove emissions in adducts formation during
the winter season. Although DNA adducts are considered to
be capable of detecting children exposure to urban air
pollutants and ETS (7), a recent paper (70) has shown that
the chemical nature of the lung DNA adducts in smokers
(measured by the nuclease P1-mediated 32P-post-labeling
assay) that appear on TL chromatograms as diagonally
radioactive zone (DRZ) is different from the typical adducts
generated by PAH or aromatic amines. Indeed, DRZ adducts
reflect exposure to aldehydes, butadiene, and cathecols, the
levels of which are orders of magnitude higher in cigarette
smoke than the carcinogenic PAHs. This raises the issue of the
validity of the interpretation of the measured DRZ adducts as
PAH-specific-smoking related DNA adducts. However, DNA
adducts in circulating lymphocytes, as measured in this study
and most of the epidemiologic researches, remain useful
markers of exposure to environmental pollutants including
those generated during the combustion process of organic
material. Higher levels of DNA adducts were detected in girls
than in boys, particularly in indoor emission exposed girls. We
have no clear explanation of the role of gender although
a higher susceptibility to PAHs contained in cigarette smoke
has been reported for females (71–73).
Cytogenetic markers were used to assess exposure of
children to airborne pollutants, soil and drinking water
contaminants. There are few studies on DNA damage in
children environmentally exposed to lead. Dalpra et al. (29)
did not find the difference in SCE frequency between 3- and
14-year-old children living in a widely contaminated area
and showing an increased absorption of lead (n ¼ 19, SCE ¼
9.0) and 5–11-year-old referents (n ¼ 12, SCE ¼ 8.9).
Bauchinger et al. (27) carried out a study in 30 children living
in a town with a lead plant. Despite a significantly increased
lead level in exposed children as compared with referent ones
no increase of chromosomal aberrations could be detected.
Yanez et al. (31) indicated increased SSB in children exposed
to arsenic (n ¼ 20, urinary arsenic ¼ 136 mg/g creatinine) and
lead (n ¼ 20, PbB ¼ 11.6 mg/dl) in the mining site as compared
with control children (n ¼ 35, urinary arsenic ¼ 34 mg/g
creatinine) and lead (n ¼ 35, PbB ¼ 8.3 mg/dl). The increased
DNA damage detected in Mexican children exposed to metals
could be explained by genotoxicity induced either by arsenic
or by lead, with a more significant effect of the latter.
Chromosome damage has been reported in adults occupationally exposed to lead (11,12,15,20,26,74). A more clear relation
than in the case of environmental exposure to lead was found
between the level of biomarkers of genetic damage and
exposure to air pollutants with clearly increased MN frequency
detected in exposed compared with unexposed children (7).
The findings were confirmed by the analysis of MN frequency
in buccal epithelium cells measured in 60 children aged
16–17 years from Katowice, the geographical area from
which our study population comes from, 60 children from
Cz˛e stochowa, and 62 children from Nowy S˛a cz, a rural area
(75). The results of the study showed a relationship between the
MN frequency and exposure to air pollutants, mainly PAHs,
with mean MN level of 9.85; 4.2 and 2.9 in Katowice,
Cz˛e stochowa, and Nowy S˛a cz, respectively. Moreover, in the
late 80s three groups of 7–15-year-old Polish children from
three areas with different level of air pollution were studied
and SCE in PBL were measured (76). The highest mean value
of SCE was observed in children from the area surrounding
a metallurgical plant (10.7), compared with lower levels
detected in children from central Krakow (7.4) and from a
rural area (6.5).
The mean level of MN found in our study was 4.44 per
1000 binucleated cells, a mean value that is in line with that
reported as a baseline value estimated by meta-analysis based
on the results of 440 subjects (4.48) (77). However, although
total MN level was similar to the expected baseline, it was
significantly higher (6.4) in children with PbB level 10 mg/dl,
the current CDC threshold of concern for Pb exposure (78).
The contribution of lead exposure in MN induction is also
suggested by the significant correlation between PbB concentration and MN level in PBL (Figure 1). The increased level of
MN in boys may be also related to higher concentrations of
PbB in boys than in girls. The observed increase of MN level
with age is in accordance with the findings obtained from a
pooled analysis showing a clear positive association of MN
frequency with age (77). Although the association between
301
D.Miel_zyńska et al.
PbB and MN levels was revealed by multiple regression
analysis, this finding needs to be confirmed by future
studies given the non chemical-specificity of this cytogenetic
biomarker.
The mean number of SCE per cell detected in our children
(7.87) was similar to that reported for children living in central
Krakow (7.4) (67), and lower than in children exposed to
ETS in the USA (10.07) (79). The positive significant
correlation between DNA adducts and SCE suggests that
increased level of SCE can be related to the exposure to
environmental agents that are capable to react with DNA. The
frequency of SCE was significantly associated with gender,
with more SCE scored in girls than in boys. This is likely to be
associated with a higher level of DNA adducts in girls.
Although SCE in children have not been consistently reported
to reflect specific environmental exposures (7), a high
frequency of SCE was found in children living downwind
of a chemical disposal site (80). A significantly increased level
of SCE was detected in our study population among children
who were probably exposed to ambient PAHs as we can
conclude from the increased level of DNA adducts in these
children. The association between DNA adducts and SCE
was also detected in multiple regression analysis although it
failed to reach statistical significance.
Our findings are suggestive of a potential role of exposure to
environmental agents, namely Pb and PAHs, in the occurrence
of cytogenetic effects as MN and SCE in peripheral lymphocytes of children. It is confirmed by the observed relationship
between DNA adducts and SCE as well as between PbB
and MN (all measured in blood). The study indicated that
there are at least two main sources of PAHs and Pb contributing
to environmental pollution: outdoor and indoor and that
emissions from coal stoves are important contributors to the
total exposure to PAHs and Pb in our region. Environmental
tobacco smoke seems to be more important as the source of
lead than of PAHs. However, because of the limited sample
size of our study, the non-exposure specific nature of the
cytogenetic markers and their high variability, and the fact that
the measured biomarkers may reflect exposures that occurred
shortly before blood sampling (e.g. PbB, 1-OHP and DNA
adducts) or a persistent/cumulative genotoxic effect (e.g. MN
and SCE), our findings need to be confirmed by other
independently conducted research.
Acknowledgements
This work was supported by the Polish Committee of Scientific Research
(3 P05D 052 24) and by the European Concerted Action CHILDRENGENONETWORK (QLK4-CT-2002-02198).
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Received on March 6, 2006; revised on July 3, 2006; accepted on July 7, 2006
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