Children and the unborn child

5. Specific substances, examples of exposure and effects

5.1 Alcohol
5.2 Tobacco smoke
5.3 Ambient air pollution
5.4 Pesticides
5.4.1 Organophosphates
5.4.2 Carbamates
5.4.3 Lindane
5.4.4 Paraquat
5.5 Drugs
5.5.1 Chloramphenicol
5.5.2 Sulfonamides
5.5.3 Diethylstilboestrol (DES)
5.6 Polychlorinated biphenyls (PCBs)
5.7 Polychlorinated dibenzo-p-dioxins (PCDDs)
5.8 Polybrominated diphenyl ethers (PBDEs)
5.9 Phthalates
5.10 Lead
5.11 Mercury
5.12 Copper
5.13 Boric acid
5.14 Nitrate and nitrite
5.15 References


This section will illustrate some differences in biological susceptibility and/or exposure between children and adults by giving a brief overview of selected chemical substances for which such differences have been reported in the literature. It should be stressed that the section is only intended for presenting some illustrative cases and therefore by no means should be considered as an exhaustive overview covering all cases and aspects.

5.1 Alcohol

Consumption of alcohol during pregnancy is associated with a great potential for developmental defects in the unborn child.
Alcohol is distributed with body water, and adverse effects in most organs have been reported. The critical target organs are the central nervous system and the liver. (Andrews & Snyder 1991, ILSI 1999).

A distinct dysmorphic condition, the foetal alcohol syndrome (FAS), has been associated with alcoholism in the pregnant mother. The abnormalities most typically associated with FAS include central nervous system dysfunction characterised by mental deficiency and microcephaly, growth deficiencies (both length and weight), a characteristic cluster of facial abnormalities (short palpebral fissures, hypoplastic upper lip with thinned vermilion, diminished or absent philtrum, deficient eye growth, short nose), and variable major and minor malformations (cardio-vascular and skeletal defects).
The critical gestational stage at which the foetus is most vulnerable to the effects of alcohol is not yet known; however, it may be about the time of conception or early pregnancy, when heavy maternal drinking is particularly risky.
The severity appears to be related to the extent of alcohol consumption by the mother during pregnancy. Data are insufficient to state a threshold and no level of maternal drinking can be established as absolutely safe for the foetus. The National Board of Health in Denmark (personal communication), recommends that pregnant women generally should avoid drinking alcohol and maximum 1 glass a day corresponding to around 12 g of alcohol per day.
The pathogenesis of FAS still remains undefined. However, three main, nonexclusive mechanism that may explain the genesis of FAS include impaired placental/foetal blood flow, deranged prostaglandin balance, and direct effects of alcohol (or acetaldehyde, the primary metabolite) on cellular processes.
(Schardein 2000a, Andrews & Snyder 1991, ILSI 1999).

Many studies in laboratory animals clearly demonstrate that alcohol is a potent developmental toxicant inducing growth deficiency, mortality, and malformations similar to those seen in humans. Acetaldehyde, the primary metabolite of ethanol, is teratogenic in mice and induces growth retardation and malformations in rat foetuses similar to those of ethanol when injected into the maternal animal during organogenesis. (Schardein 2000a).
Animal models have provided evidence that foetal alcohol effects are dose-dependent and have clearly demonstrated that the peak maternal blood alcohol concentration and the pattern of drinking are the most important determinants of the magnitude of alcohol-related birth defects. Different aspects of development may be sensitive to different levels of alcohol. (ILSI 1999).

5.2 Tobacco smoke

Suspicions that exposure to tobacco smoke could be hazardous to reproductive function and foetal development date back to early in the present century. It is now generally considered that smoking during pregnancy increases the risk of perinatal mortality, lowers mean birth weight, increases the risk of spontaneous abortion, and has a significant influence of risks of premature delivery, placenta previa, and abruptio placentae. However, it appears that congenital malformations are not associated with smoking. (Schardein 2000b).

Exposure to environmental tobacco smoke (passive smoking) has been reported to be associated with a variety of respiratory disorders in children, including asthma, wheezing, dyspnoea, cough, bronchitis, pneumonia, otitis media, and laryngospasm with anaesthesia induction. Children exposed to passive smoke in the home have more days of school absence, bed confinement, and restricted activity than children in a smoke-free environment. (McCance & Huether 1998).

Tobacco smoke contains more than 3800 different compounds. About 10% of these constitute the particulate phase, which contains nicotine and tar. The remaining 90% contains volatile substances such as carbon monoxide, carbon dioxide, cyanides, various hydrocarbons, aldehydes, and organic acids. Although all of these substances affect the smoker to some degree, nicotine is generally considered to be the primary substance responsible for the pharmacological responses to smoking.
The nicotine content is approximately 3 mg/cigarette smoked in 5 to 10 minutes. Pharmacologically, it is a classic cholinergic agonist initiating nervous stimulation and then blocking cholinergic receptors. By releasing catecholamines, nicotine increases heart rate, oxygen consumption, utilisation of free fatty acids, hypoglycaemia, all of which can be expected to affect foetal development. Nicotine freely crosses the placenta but levels in foetal tissues remain relatively low compared with those of maternal tissues, except in the early stages of development.
(Schardein 1993b, IARC 1986).

Smoking reduces birth weight of offspring and the consensus of over 200 published studies is that smokers’ babies weigh, on the average, 170 to 200 g less at birth than non-smokers’ babies, with about twice as many babies weighing less than 2500 g at birth. Reduced birth weight is primarily due to intrauterine growth retardation, rather than prematurity. The cause of the growth retardation in utero remains highly controversial, but the available published information supports the direct effects of nicotine and carbon monoxide as factors causing intrauterine hypoxia as the most likely mechanism for the effect of smoking on birth weight. There seems to be a dose-response relation as it has been calculated that the decrease in birth weight is about 8 to 9 g for each cigarette smoked daily. Smoking fewer than seven to ten cigarettes per day or early cessation of smoking in pregnancy results in foetal body weights that do not differ significantly from those of nonsmokers.
Increased perinatal or neonatal mortality is also associated with smoking during pregnancy. Perinatal mortality increases 20% for less than one-pack-per-day and 35% for more than one-pack-per-day and in some studies, the risk is more than doubled. The incidence is proportional to the birth weight. Prematurity, anoxia, and placental complications (abruptio placentae and placenta previa) are generally regarded as being responsible for most of this increase.
Spontaneous abortion rates may also be higher in mothers who smoke. The frequency of abortion appears to be directly related to the number of cigarettes smoked with the risk perhaps twofold greater among heavy (more than one pack per day) smokers than in nonsmokers. Heavy smokers also appear to abort earlier in pregnancy.
Data are conflicting concerning a possible causal association between congenital malformations and cigarette smoking. Of 21 studies published since 1970, 11 studies reported positive associations for congenital malformations; cardiac defects, cleft lip and palate, and anencephaly were most often identified as specific malformations associated with smoking. The remaining 10 studies failed to associate smoking in pregnancy and malformations.
The effects of maternal smoking on functional development have not been defined precisely but several reports suggest some causal association between smoking and impaired function. New-borns of smoking mothers have been reported to perform poorer in two operant tasks (head turning and sucking), to be less visually alert, and to have atypical sleep patterns. Childhood hyperkinesis is reportedly more common among children of smoking mothers. Maternal smoking may also have an adverse effect on learning. Recent studies have shown that these types of effects are not limited to offspring of active smokers alone.
(Schardein 2000b).

Some of the developmentally toxic effects observed in humans have been observed in laboratory animals as well. Foetal growth retardation is also a characteristic in rats and rabbits, in the absence of malformations, at human exposure levels. Increased stillbirths have also been observed in rabbits exposed to the equivalent of 20 cigarettes per day during pregnancy. (Schardein 2000b).

5.3 Ambient air pollution

In the past decades the incidence of respiratory allergic diseases in children has increased significantly in the Western world. Current data do not indicate that ambient air pollution attribute significantly to this rise, however, air pollution may cause significant increases in respiratory symptoms and their intensity in persons suffering from asthma and other respiratory diseases.

Children in this respect may be considered as a special risk group as they breathe more air relative to their body weight and lung surface, and thus receive proportionally higher doses of air pollutants systemically and locally. Furthermore children spend more time outdoor, and are more active, resulting in mouth-breathing and increased respiratory rate (see section 2).

Recently Jedrychowski et al. (1999) in a Polish study followed groups of children in two different areas, and over time compared their growth in lung function. They found a significantly greater growth in lung function among the boys from the area with low level air pollution compared to boys from the area with the higher air pollution level. For girls a similar difference did not reach the level of significance. The authors suggested that air pollution may lead to retardation in pulmonary function growth during the pre-adolescent years.

Several epidemiological studies indicate that children are susceptible to ambient air pollution. Increasing levels of ambient ozone levels have been shown to be associated with a decrease in lung function in children (Krzyzanowski et al. 1992, Hoek et al. 1993a, Hoek et al. 1993b, Braun-Fahrländer et al. 1994, Stern et al. 1994, Kinney et al. 1996).

In the study by Krzyzanowski et al. (1992), children together with asthmatics and adults spending long time outdoor were identified as having the most significant decline in lung function in relation to ozone exposure.

Burnett et al. (1994) studied the association between ambient air pollution levels and the number of daily hospital admissions at 168 hospitals in the Ontario area, representing 8.1 million people. Summer respiratory admissions were found to be closely related to ozone levels, and children was found to be especially affected as 15% of the admissions for children were associated with ambient air pollution compared with 4% for the elderly population.

Recently Loomis et al. (1999) and Woodruff et al. (1997) found increased infant mortality associated with ambient levels of fine particles. Loomis et al. (1999) found that a 10 m g/m3 increase in mean level of fine particles during a period of three days was associated with an 6.9% increase in infant mortality. The authors noted that this particle-related excess mortality observed among children was greater in relative risk terms than excess mortality observed among elderly people in several other studies.

In relation to average particle level (long term exposure) Woodruff et al. (1999) compared post-neonatal mortality in areas with different levels of air pollution. They found a significantly increased odds ratio of 1.10 for postneonatal mortality (adjusted for other covariates) in highly polluted areas compared to low pollution areas (mean particle levels of 44.5 g/m3 and 23.6 g/m3, respectively). The increased mortality was primarily due to respiratory-related mortality and sudden infant death syndrome. The authors suggested that children may be considered as a susceptible group in relation to particulate air pollution.

5.4 Pesticides

5.4.1 Organophosphates

Organophosphorus insecticides are normally esters, amides, or thiol derivatives of phosphoric, phosphonic, phosphorothioic, or phosphonothioic acids. More than 100 different organophosphorus insecticides are known. Organophosphorus insecticides exert their acute effects in both insects and mammals by inhibiting acetylcholinesterase (AChE) in the nervous system with subsequent accumulation of toxic levels of acetylcholine (ACh), which is a neurotransmitter. Delayed neuropathy is initiated by attack on a nervous tissue esterase distinct from AChE. (WHO 1986a).

Many thousands of cases of acute poisoning by organophosphorus insecticides have been recorded, the majority being due to parathion and methyl parathion (WHO 1986a).
Great differences in response to parathion exposure have been observed among individuals but children are generally more sensitive than adults.
In a number of cases, the oral dose of parathion was known to be exactly 900 mg (around 15 mg/kg b.w.), and it was uniformly fatal. However, in one case the ingestion of 120 mg parathion (around 2 mg/kg b.w.) led rapidly to the death of a man. On the other hand, children (5-6 years old) died after eating 2 mg of parathion (around 0.1 mg/kg b.w.) In instances in which parathion-contaminated food was eaten by people of different ages, death occurred mainly or exclusively among children. Also the epidemiological evidence indicates that parathion is generally more toxic to children than to adults. (Gallo & Lawryk 1991).

The acute toxicity of different organophosphates ranges from highly toxic to only slightly toxic (oral LD50-values for the rat range from less than 1 mg/kg b.w. to over 3000 mg/kg b.w.) (WHO 1986a). Several studies have reported higher sensitivity based on lethality in young animals compared to adults following acute exposure to organophosphorus insecticides. The age-related differences in sensitivity may differ qualitatively and quantitatively with different organophosphates and varying exposure conditions (e.g., high vs. low dose, acute vs. repeated). A study by Liu et al. (1999) showed that repeated exposures to chlorpyrifos were associated with relatively similar degrees of cholinesterase inhibition among the age groups (neonatal, 7 days of age vs. adult, 90 days of age). In contrast, cholinesterase activity and muscarinic receptor binding were generally more reduced in neonatal relative to adult brain regions following repeated exposures to methyl parathion.

Many organophosphorus insecticides are embryotoxic at doses that are toxic for the mother, but only few teratogenic effects have been reported (WHO 1986a).
Parathion is toxic to the foetus causing prenatal and postnatal death of the young and reduced weight gain of the surviving young; no developmental abnormalities have been observed. The greater susceptibility of the foetus is in spite of the fact that inhibition of cholinesterase activity in foetal blood is less than that in maternal blood following dosing of the mother. Weanlings are also more susceptible than adults to parathion due to their poorly developed microsomal enzymes and also greater inherent susceptibility of the young brain. (Gallo & Lawryk 1991).

In reproduction studies of methyl parathion at maternally toxic dose levels (ChE inhibition), no consistent effects on litter size, number of litters, pup survival rates, and lactation performance were observed; no primary teratogenic or embryotoxic effects were noted. (WHO 1993b).
In a toxicokinetic study in rats orally exposed to chlorpyrifos during late gestation (gestational days 14 to 18), the concentration of 3,5,6-trichloro-2-pyridinol (TCP - the only metabolite detected) in the foetal brain was two- to fourfold higher than the concentration in the maternal brain. The concentration of TCP in the maternal liver was approximately fivefold higher than in the foetal liver. Thus, the foetal nervous system may be exposed to a higher concentration of chlorpyrifos than the maternal nervous system when the dam is exposed during late gestation.
Cholinesterase activity levels were determined in liver and brain from maternal and foetal rats. Maternal liver cholinesterase was two-fold more inhibited than foetal liver activity. Similarly, the maternal brain exhibited more cholinesterase inhibition than foetal brain at 5 hours after the last dose of chlorpyrifos. Foetal brain cholinesterase activity recovered to control levels by 24 hours after the last dose, whereas the maternal brain activity was still inhibited (44%) 5 days after the last dose.
(Hunter et al. 1999).

In USA, chlorpyrifos is one of the most commonly used pesticides in the indoor environment today. A recent study (Gurunathan et al. 1998) showed that after a single broadcast spraying of chlorpyrifos in the indoor environment, chlorpyrifos continued to accumulate on children’s toys and hard surfaces 2 weeks after spraying. Based on this study and other research studies, the estimated chlorpyrifos exposure levels for children from indoor spraying are approximately 20 to 120 times above the current recommended reference dose (US-EPA) of 3 g/kg b.w./day for chlorpyrifos exposure to children from all sources. (Davis & Ahmed 1998). In Denmark, pesticides are not used very much in the indoor environment (MST 2000 - personal communication).

5.4.2 Carbamates

A considerable number of reproduction and teratogenicity studies have been carried out with different carbamates and various animal species. Generally, the foetal effects included an increase in mortality, decreased weight gain in the first weeks after birth, and induction of early embryonic death. Certain carbamates also induce teratogenic effects, mainly at high dose levels applied by stomach tube. When the same dose levels was administered with the diet, no effects were seen. (WHO 1986b).

5.4.3 Lindane

Lindane is the g -isomer of 1,2,3,4,5,6-hexachlorocyclohexane (HCH). Lindane interacts with cellular membranes and may produce several generalised cytotoxic effects associated with impaired membrane function. (ATSDR 1997).
Several cases of fatal poisoning and of non-fatal illness caused by lindane have been reported. The toxic or lethal dose appears to vary considerably; under certain conditions, 10 to 20 mg/kg b.w. can represent a lethal hazard to humans, but higher doses can be tolerated when followed by appropriate medication. Some information indicates that children are more sensitive to lindane than adults. (WHO 1991b).
The acute toxicity in experimental animals is moderate (oral LD50-values for rats and mice range from 60 to 250 mg/kg b.w. depending on the vehicle used). Young animals are generally more sensitive than adults. (WHO 1991b).
Lindane had no teratogenic effect in experimental animals after oral or parenteral administration. Foetotoxic and/or maternal toxic effects were observed with doses of 10 mg/kg b.w. and above when given by oral gavage. (WHO 1991b).

5.4.4 Paraquat

Paraquat is the dichloride salt of 1,1’-dimethyl-4,4’-bipyridinium ion.
Following exposure to paraquat, a characteristic lung injury is observed.
A large number of cases of suicidal or accidental poisoning from paraquat has been reported. Among individuals who experience lung damage the mortality rate is high. The minimum lethal dose is stated to be about 35 mg/kg b.w.; however, some individuals have survived after ingesting 10-20 g of paraquat (about 140-280 mg/kg b.w.).
The acute toxicity in experimental animals is moderate (oral LD50-values for rats range from 100 to 200 mg/kg b.w.). Young rats were more resistant to the toxicity of paraquat than older rats; and some authors have paralleled this resistance with that of young rats to oxygen toxicity. One study (Smith & Rose 1977b) found a more than 40% increase in cumulative mortality in 180 g rats compared with 50 g rats, after oral dosing with paraquat at 175 mg/kg b.w.; according to the authors, the difference in renal function between young and mature rats accounted for the difference in paraquat toxicity.
Oral administration of high doses of paraquat to pregnant rats, mice, or rabbits on various days of gestation produced no evidence of teratogenicity, but did produce slight embryotoxicity; however, the doses used induced significant maternal toxicity.
(WHO 1984).

5.5 Drugs

Several drugs are known to induce various adverse effects in children including the unborn child. The classical example is the known human teratogen thalidomide which caused unusual limb defects in babies born of mothers who had taken thalidomide between the fourth and ninth weeks of pregnancy. A few other examples are given here.

5.5.1 Chloramphenicol

Chloramphenicol is a broad-spectrum antibiotic having an antibacterial spectrum and potency very similar to those of the tetracyclines. It is well absorbed from the gastrointestinal tract, metabolised in the liver, and excreted rapidly in the urine predominantly as metabolites. Chloramphenicol crosses the placenta and the concentration in the foetus varies from 30 to 80% of the concentration in maternal blood. Chloramphenicol is also secreted into the milk.
Premature babies and infants (up to one month of age) do not have the detoxifying liver enzyme (glucuronyltransferase) and are therefore extremely sensitive to the serious toxic effects associated with exposure to chloramphenicol. The symptom complex observed in this age group, often referred to as the "grey baby syndrome", consist of gastrointestinal effects (vomiting and abdominal distension), dyspnoea, cyanosis, and vascular collapse.
Because of the serious toxicity such as aplastic anaemia (complete suppression of bone-marrow activity) observed in adults and in foetuses and the "grey baby syndrome" in infants following systemic administration, chloramphenicol is only indicated for the treatment of very serious infections; it is contraindicated for pregnant and lactating women. No data on teratogenic effects in humans are available.
(Lægemiddelkataloget 1996a, Goth 1978).

Chloramphenicol caused closure defects, among other abnormalities, in rats and non-specific malformations in rabbits. The teratogenic activity in the rat was attributed to interference with activity of the electron transport systems and oxidative energy formation in the embryo during embryogenesis. In mice and rhesus monkeys, no teratogenic effects have been noted under the regimens employed. Some postnatal behavioural effects, including reduction in learning ability, have been described in mice following prenatal treatment. (Schardein 2000c).

5.5.2 Sulfonamides

Sulfonamides are derivatives of sulfanilic acid and used for treatment of infections predominantly of the urinary tract. Only one sulfonamide, sulfamethizole, is registered in Denmark.
Sulfonamides are rapidly and extensively absorbed following oral administration, metabolised (acetylation or glucuronidation) in the liver and excreted in the urine predominantly as metabolites. Sulfonamides cross the placenta and are secreted into milk.
The most common adverse effects (dermatitis and drug fever) of sulfonamides are related to acquired hypersensitivity. Serious toxic effects (icterus because of damaged liver cells, aplastic anaemia, neutrocytopenia, acute haemolytic anaemia) are infrequently reported.
Because of the risk of development of icterus in new-borns and young infants (age up to one month), sulfonamides are contraindicated to pregnant women during the last 4 weeks before expected delivery, to lactating women, and to infants (age up to one month).
(Lægemiddelkataloget 1996b).

One study of 458 women taking sulfonamides over the entire pregnancy reported that there were more congenital malformations among their offspring than in the young of untreated controls. Four other studies found no relation to sulfonamide therapy in early pregnancy and malformations. A large collaborative study found no significant malformations associated with the use of specific sulfonamides, including sulfamethizole.
Sulfonamides have shown a mixed teratogenic potential in experimental animals with about one-third indicating activity.
(Schardein 2000d).

5.5.3 Diethylstilboestrol (DES)

DES is an artificial non-steroid oestrogen used as an antineoplastic agent for treatment of prostate cancer. DES is absorbed rapidly following oral administration, metabolised in the liver and excreted in the urine and faeces. (Lægemiddelkataloget 1996c, Schardein 2000e).

DES is a transplacental carcinogen in humans and development of adenocarcinomas in the cervix and vagina have been observed in young females of mothers treated with DES during pregnancy.
DES was apparently thought to be efficacious in the definitive and preventive treatment of abortion and premature delivery. For this reason, it was given to a large number of pregnant women, being approved for use in pregnancy by the US-FDA in 1947. Early in 1970, seven cases of vaginal adnocarcinoma were reported in young women aged 15 to 22 and it was found that the patient’s mothers had ingested oestrogen, DES specifically, in the first trimester of their respective pregnancies many years earlier. In 1971, a registry of clear cell adenocarcinoma of the genital tract in young females was established. By June 1997, 695 cases of clear cell adenocarcinoma were listed in the registry, two-thirds of which
were associated with prenatal DES exposure. In all of the patients who have vaginal and cervical carcinoma, maternal ingestion of DES occurred before the 18th week of pregnancy and thus, early first-trimester exposure appears to be mandatory in its subsequent toxicity. A peak in the age incidence curve of DES-related cases has been observed at about 19 years, with the age range (latency) being 7 to 30 years.
By far, most of the cases reported have been in the USA, but case stories have also been reported from other countries; however, countries where DES was never used (e.g., Denmark and West Germany) did not have cancer cases.
A mechanism for the vaginal lesions has been theorised. DES may act to sensitise the proliferating stroma of the lower mullerian duct so that it is incapable of fostering upgrowth of urogenital sinus epithelium to spread over and replace the epithelium covering the vagina and cervical protico by 18 weeks when this event should occur. DES may also preferentially affcet the stroma of the developing cervix.
(Schardein 2000e).

5.6 Polychlorinated biphenyls (PCBs)

Because of their high persistence, and their other physical and chemical properties, PCBs are present in the environment all over the world. The general population is exposed to PCBs mainly through contaminated food (aquatic organisms, meat, and dairy products). Infants are exposed through the mother’s milk and it has been estimated that the nursing period contributes about 1.3% of the life-time intake. (WHO 1993a).
In Denmark, an investigation has been performed to determine the content of dioxins and PCBs in the breast milk of Danish mothers. The average concentration of total PCB was 469 ng/g fat corresponding to an average daily intake in the sucking child of 2.4 g/kg b.w./day. In the report, a NOAEL of 0.33 g/kg b.w./day was considered for behavioural effects. However, these two figures should not be compared as the TDI relates to a life long exposure. (SST/FDIR 1999).

In general, PCBs appear to be rapidly absorbed, particularly via the gastrointestinal tract after oral exposure; information on the rates of human absorption is limited. PCBs are rapidly cleared from the blood and accumulate in the liver and adipose tissue. There is evidence of placental transport, foetal accumulation, and distribution to milk. (WHO 1993a).

There are great difficulties in assessing human health effects separately for PCBs and polychlorinated dibenzofurans (PCDFs) since, quite frequently, PCB mixtures contain PCDFs.
Therefore, in many cases, it is not clear which effects are attributable to the PCBs themselves and which to the much more toxic PCDFs. Much of the data of the human toxicity come from large episodes of intoxication, e.g., the Yusho and Yu-Cheng episodes. The most striking signs of intoxication in Yusho and Yu-Cheng patients include hypersecretion in the eyes, pigmentation of the nails and mucous membranes, and acneiform eruptions of the skin.
Furthermore, oedema of the extremities, liver enlargement and disorders, central nervous disturbances, respiratory problems, and changes in the immune status were also observed. In children of Yusho and Yu-Cheng patients, diminished growth, dark pigmentation of the skin and mucous membranes, gingival hyperplasia, xenophthalmic oedematous eyes, dentition at birth, abnormal calcification of the skull, rocker bottom heel, and a high incidence of low birth weight were observed. (WHO 1993a).

The Michigan Maternal and Infant Study has reported adverse developmental health outcomes in new-born infants of mothers who consumed more than 12 kg of contaminated Great Lakes fish. Statistically significant decreases in infant’s birth weight, gestational age, and head circumference compared to controls were observed. The infants also exhibited neurodevelopmental and behavioural deficits based on tests of visual recognition and memory at 7 months and 4 years of age. At 11 years of age, many of the neurobehavioural deficits had persisted, e.g., poorer short- and long-term memory and lower IQ scores. (Johnson et al. 1999).

Lesions induced in experimental animals exposed to PCB mixtures or individual congeners concern the liver, skin, immune system, reproductive system, oedema, and disturbances of the gastrointestinal tract and thyroid gland.
The acute toxicity of PCBs is generally low in rats; young animals appear to be more sensitive than adults.
The Rhesus monkey is the most sensitive species with regard to general toxicity and particularly with regard to reproductive toxicity. PCBs adversely affected the reproductive performance of female Rhesus monkeys, mated with control males after 6 months of dietary exposure to a toxic dose (0.09 mg/kg b.w./day) and continuation of the exposure for an average of up to 10 months. Neonates of nursing mothers exposed to PCBs (during gestation and lactation of the first generation) showed adverse effects similar to those seen in their mothers and, in addition, persistent behavioural disturbances as well as several other adverse effects. PCBs may also bind to the cytoplasmic oestrogen receptor. Effects have also been observed on the oestrus cycle of female rats and monkeys, on the sex organs of male rats, and on the implantation rate of fertilised ova following exposure of female mice or male rats.
Available studies in rats and monkeys did not indicate any teratogenic effects (malformations), when animals were dosed orally during organogenesis at doses that produced foetotoxicity and/or maternal toxicity.
(WHO 1993a).

5.7 Polychlorinated dibenzo-p-dioxins (PCDDs)

Consumption of food (including human milk) is the most important pathway for exposure to PCDDs for the general population (adults and children) representing over 90% of the total daily intake. Other pathways include inhalation and direct contact with PCDDs. Exposure of infants and young children may be very high because of their relatively high consumption of milk, including breast milk. (ATSDR 1998).
In Denmark, an investigation has been performed to determine the content of dioxins and PCB in the breast milk of Danish mothers. The average concentration of dioxins + PCB was 30 pg TEQ/g fat (TEQ = 2,3,7,8-tetrachlorodibenzo-p-dioxin toxicity equivalents) corresponding to an average daily intake of the sucking child of 150 pg TEQ/kg b.w./ day. A TDI of 1-4 pg TEQ/kg b.w./day (dioxins and dioxin-like PCBs) has been set (WHO 1999) as an average daily intake fo the whole lifetime (SST/FDIR 1999).

Humans can absorb PCDDs by the inhalation, oral, and dermal routes of exposure; absorption is vehicle-dependent and congener-specific. For most mammalian species, the liver and adipose tissue are the major storage sites of PCDDs. Tissue deposition is congener-specific and depends on the dose, the route of administration, and age; 2,3,7,8-substituted PCDDs are the predominant congeners retained in tissues. PCDDs are very slowly metabolised. The major routes of excretion are the bile and the faeces; smaller amounts are excreted via the urine. In mammalian species, lactation is an effective way of elimination PCDDs from the liver and other tissues. Human studies show that infants may absorb up to 95% of the amount ingested via breast milk. (ATSDR 1998).

A wide variety of effects have been observed in adults exposed to 2,3,7,8-TCDD. The primary targets appear to be the skin, liver, thyroid, and cardiovascular, endocrine, and immune systems; an increased cancer risk has also been observed. It is likely that these organs/systems will also be sensitive targets in children. Children exposed to 2,3,7,8-TCDD appear to be more sensitive than adults to the dermal effects (chloracne). A number of human studies have investigated the potential of 2,3,7,8-TCDD to induce developmental effects. In one study, no significant increases in the incidence of birth defects have been observed in the children of parents living in Seveso at the time of the accident or during the next 6-year period. In contrast, other studies have found increases in specific types of defects, although the total number of defects was not significantly altered. (ATSDR 1998).
In animal oral toxicity studies, toxic effects of PCDDs have been observed in most organs/systems. The studies clearly demonstrate that the developing organism is very susceptible to the toxicity of PCDDs, in particular 2,3,7,8-TCDD. Prenatal or perinatal exposure has resulted in structural malformations (e.g., cleft palate, hydronephrosis), functional alterations (e.g., damage to the immune system, impaired development of the reproductive system), decreased growth, and foetal/new-born mortality in several animal species. The organ system most sensitive during development is the reproductive system (alterations in androgen levels, secondary sex organs, spermatogenesis, fertility, and sexual behaviours). Also neurobehavioral effects are observed. Additionally, several animal studies provide evidence that exposure to 2,3,7,8-TCDD via mother’s milk alone can adversely affect the developing animal. (ATSDR 1998).

5.8 Polybrominated diphenyl ethers (PBDEs)

PBDEs are widely used in a variety of materials and products including textiles, many types of electronic devices, cabins for and circuit boards in personal computers and TV sets, electrical cables, switches and capacitors, and building materials. Furthermore, PBDEs are found in several foods of animal origin (fish, meat, and cow’s milk) (Darnerud et al. 1998, WHO 1994, WHO 1997b).
In Swedish human milk samples, the concentration of PBDEs has increased continuously with levels showing an exponential increase from 1972 and a doubling time of 5 years (Norén & Meironyté 1998).

Studies on the reproductive toxicity of PBDEs are limited. Only one study on fertility (decaBDE in rats) is available; no treatment-related effects in reproductive performance or maturation of pups were reported.
The developmental toxicity studies available on decaBDE, octaBDE, and pentaBDE are equivocal. Foetotoxicity (resorptions, delayed ossification) but no malformations have been observed in rats exposed to decaBDE (one study) and octaBDE (two studies) and in rabbits exposed to octaBDE (one study) at dose levels which did not induce maternal toxicity. In contrast to these findings, foetotoxicity was observed in rats exposed to pentaBDE (one study) only in the presence of maternal toxicity. (WHO 1994, Darnerud et al. 1998).

A very recent study investigating possible neurobehavioural effects in neonatal mice (10 days old) following a single exposure to pentaBDE suggests differences in behavioural patterns between treated and control mice (Eriksson et al. 1998).

5.9 Phthalates

Phthalates are high production volume chemicals widely used as additives in PVC plastics. Due to the ubiquitous use, phthalates are found everywhere in the environment.

In Denmark, the intake of the phthalates DEHP, DBP, and BBP has been analysed in a number of samples (29) from a double portion study where adults participated; estimates of the mean and maximum intake were 0.13-0.29, 0.02-0.03, and 0.19-0.3 mg/person/day for DBP, BBP, and DEHP, respectively. For children, the mean and maximum intake of phthalates from infant formulae has been estimated to <0.042, 0.0006-0.0009, and 0.009-0.021 mg/child/day (assumed child weight of 3 kg) for DBP, BBP, and DEHP, respectively. (Petersen 1999).

Some phthalates affect fertility and reproduction in rodents of both sexes and also produce developmental effects in the offspring.

Generally, phthalates with side chains of 4 to 6 carbons atoms in length, e.g., di-2-ethylhexyl phthalate (DEHP), dibutyl phthalate (DBP), and butyl benzyl phthalate (BBP), affect the reproductive system of male rodents whilst phthalates with side chains shorter than 4 carbon atoms and longer than 6 carbon atoms appear to be without effect. Observed effects include marked reductions in the weights of the testes and accessory sex glands, decreased numbers of spermatocytes, degeneration of the seminiferous tubules, a reduction in testicular zinc and iron levels and serum testosterone levels, an increase in testosterone levels in the testes, sloughing of germ cells, and vacuolisation of Sertoli cells (DEHP). Spe cies differences have been observed as the reproductive system of the male rat appears to be more sensitive than that of the mouse, which appears to be more sensitive than that of the hamster, guinea pig, and non-human primates.
The effects on the male reproductive system are influenced by the age at which the animal is exposed. Studies of DEHP have shown that developing and sexually immature male rats are more sensitive to DEHP-induced testicular toxicity than sexually mature rats and that the onset of the effects in young animals is more rapid. Furthermore, exposure (DEHP, DBP) of rats prenatally and during suckling has produced irreversible testicular damage at dose levels inducing only minimal effects in adult animals. (CSTEE 1998, WHO 1997, Nielsen & Larsen 1996, Mylchreest & Foster 1998).

Numerous studies have shown that some phthalates induce embryotoxic and teratogenic effects in the offspring. DEHP is embryotoxic and teratogenic in mice and embryotoxic in rats at maternally non-toxic dose levels (CSTEE 1998, Nielsen & Larsen 1996). DBP generally induce foetotoxic effects in rats and mice in the absence of maternal toxicity, and teratogenic effects only at high maternally toxic doses (WHO 1997a).
In very recent studies in rats exposed to DBP during the prenatal and early neonatal periods, a number of effects were seen in the male offspring, including decreased anogenital distance, absent or underdeveloped epididymis and seminal vesicles, cleft phallus (hypospadias), decreased reproductive organ weights, and widespread germ cell loss in the testis. In contrast, vaginal opening, age at first oestrus, and oestrous cyclicity were not affected in the female offspring indicating that DBP is not oestrogenic but rather antiandrogenic. (Mylchreest & Foster 1998).

5.10 Lead

In the non-smoking adult general population, the major exposure pathway for lead is from food and water. The level of dietary exposure to lead depends upon many factors, including foodstuffs consumed, processing technology, use of lead solder, lead levels in water, and use of lead-glazed ceramics.
For infants and young children, food, air, water and dust/soil are the major potential exposure pathways. For infants up to 4 or 5 months of age, air, breast milk, formulae and water are the significant sources of lead exposure. For infants and young children, lead in dust and soil often constitutes a major exposure pathway. Lead levels in dust depend upon such factors as the age and condition of housing, the use of lead-based paints, lead in petrol and urban density. The intake of lead will be influenced by the age and behavioural characteristics of the child and the bioavailability of lead in the source material. (WHO 1995).
A recent Swedish study (Berglund et al. in press - quoted from CSTEE 2000) has shown that food is now the main source of lead exposure even in young children living in areas with high soil lead concentrations, i.e. downtown Stockholm (<10-330 mg/kg in soil) and mining areas (20-5000 mg/kg). It was concluded that lead in soil and dust contributed little to the total intake of lead.

In humans, lead adversely affects several organ systems and organs, including the nervous, haematopoietic, reproductive, and cardiovascular systems, the liver, the kidney, and the gastrointestinal tract. Neurodevelopmental effects and subcellular changes, particularly the effects on haem synthesis, appear to be the most sensitive endpoints. (WHO 1995).
Cognitive and sensory motor deficits have been shown in children to be associated with blood lead levels as low as 100 to 150 g/l. An average IQ decrement between 1 to 3 points with increasing blood lead levels from 100 to 200 g/l correspons to 20% or less of the standard deviation of a typical IQ distribution. These and more recent data indicate, that even below 100 g/l effects might occur and that no clear threshold for effects has been identified. (CSTEE 2000).

Children, in comparison with adults, are more susceptible to lead in several respects. Children have a greater absorption of ingested lead than adults, resulting in a higher body burden from a given external exposure. About 40 to 50% of dietary lead is absorbed from the gastro-intestinal tract in infants and young children compared to around 5 to 10% in adults. Absorption of lead from ingested dust and soil is somewhat lower than from food, approximately 30% in infants and young children. It also appears that children are generally more sensitive to the toxicological effects of lead at a given internal exposure level (measured as the blood lead level) as the lowest observed effect levels (LOAELs) for various end-points (e.g. slowed nerve conduction velocity, impaired neurobehavioural function, encephalopathy, anaemia, reduced haemoglobin) are lower in children than in adults. (WHO 1995, WHO 1996).

5.11 Mercury

The general population is primarily exposed to inorganic mercury and methyl mercury through the diet. In most foodstuffs, mercury is largely in the inorganic form. Fish and fish products are the dominant source of methyl mercury in the diet. Air and water can also contribute significantly to the total daily intake of total mercury. Furthermore, dental amalgam may also be a source of exposure to inorganic mercury due to release from amalgam restorations. (WHO 1990, WHO 1991a).
No specific information was found concerning exposure of infants and children.

Methyl mercury is a well-established neurotoxicant that can cause serious adverse effects on the development and function of the human central nervous system, especially when exposure occurs prenatally (Harada 1995). The neurotoxic potential was first described from industrial exposures as the Hunter-Russell syndrome, and then reappeared in the fishing town of Minamata, Japan, in the early 1950s (Igata 1993). Most surprisingly, while unaffected themselves by mercury toxicity, many pregnant women exposed to mercury-contaminated fish bore infants that suffered from severe congenital poisoning (Harada 1995, Igata 1993). The characteristics of this form of developmental neurotoxicity are now relatively well known at high exposure levels, where a cerebral palsy syndrome occurs. In less severe poisoning, blindness, deafness, and mental retardation may be apparent. In a poisoning incident in Iraq, a dose-response relationship was established between maternal hair-mercury concentrations during pregnancy and the prevalence of severe psychomotor retardation in the children (Marsh et al. 1990). This evidence from poisoning outbreaks clearly documents the hypersusceptibility of the developing nervous system with regard to this neurotoxicant.

Current concerns relate to the neurotoxic risks at lower exposure levels prevalent in fishing communities (Grandjean et al. 1997). A birth cohort of 1000 Faroese children was examined at age 7 years, where clinical examination did not reveal any clear-cut abnormalities associated with the cord-blood mercury concentrations. However, mercury-related neuropsychological deficits at this age occurred in the domains of language, attention, and memory, and to a lesser extent in visuospatial and motor functions. The associations could not to be explained by various possible confounders such as polychlorinated biphenyls (PCBs) from seafood, and they remained after exclusion of highly-exposed children with a maternal hair-mercury concentrations above 10 g/g. This limit was thought to represent an upper safe level as based on the data from Iraq (WHO 1990).

Supporting evidence has now emerged from a fishing community in Madeira, where 149 children from the first grade in school showed mercury-related delays in the electrical signals of the brain, as recorded by the evoked potentials technique (Murata et al. 1999). A similar pattern was seen in the Faeroes (Grandjean et al. 1997). Other cross-sectional studies in Brazil (Grandjean et al. 1999) and French Guyana (Cordier et al. 1999), have shown mercury-associated developmental effects in agreement with the Faroese findings. However, a prospective study in the Seychelles has not revealed any clear adverse effects related to maternal hair-mercury concentrations, but results beyond 5 years of age are not yet available (Davidson et al. 1998).

Although the question as to the safety of fish consumption during pregnancy has not been settled, the preponderance of evidence indicates that the unborn child is much more susceptible to methyl mercury neurotoxicity than adults are. This substantial age-dependency may be a more general phenomenon for neurotoxicants, as similar evidence is available for other substances, especially lead and PCBs (Steuerwald et al. 2000).

5.12 Copper

In the general population, the major route of exposure to copper is oral. Variations in dietary copper intake reflects different dietary habits as well as different agricultural and food processing practices used world-wide. In some cases, drinking water may make a substantial additional contribution to the total daily intake of copper, particularly in households where corrosive waters have stood in copper pipes. All other intake for copper (inhalation and dermal) are insignificant in comparison to the oral route. Women using copper-containing intra uterine devices (IUDs) are exposed to only minor amounts from this source. (WHO 1998a).

Copper is an essential element and adverse health effects are related to deficiency as well as to excess intake. The relationship between intake and risk has a U-shaped curved, with risk for deficiency associated with low intakes and risk for toxicity associated with high intakes. The range of acceptable intakes which meet the biological requirement without causing toxicity may be rather narrow.
New-borns appear to absorb copper more readily than adults whereas the aged appear to absorb copper less efficiently. Copper can cross the placental barrier and is taken up by the foetus. The copper concentration in the liver of new-borns is 6-10 times higher than in adults but decreases during the first 3 months of life.
The concentration of copper in the body is kept relatively constant by homeostatic mechanisms, and toxic effects of long-term ingestion of excess copper are not frequently observed in the general population. However, children below one year of age are probably more susceptible than adults to copper toxicity because the homeostatic mechanisms may not have fully developed. A number of cases (some of them fatal) in which development of liver damage (early childhood cirrhosis) has been related to excess intake of copper from drinking water has been described.
(Nielsen 1997).

5.13 Boric acid

The most frequent and appreciable general population exposures to boron are likely to be from ingestion of food and, to a lesser extent, from ingestion of drinking water. Other potential sources include absorption of boron from cosmetic and medical preparations through mucous membranes or damaged skin; and the inhalation, dermal absorption, or accidental ingestion of boron-containing household cleaning products, pesticides, or fertilisers. (WHO 1998b).

Boric acid is readily absorbed from the gastrointestinal and respiratory tracts. The absorption is essentially complete (approximately 95% in humans) following ingestion. Dermal absorption across intact skin is negligible in all species evaluated, including humans (infants and adults). Boric acid is rapidly excreted mainly (95%) by the kidneys. (WHO 1998b).

Only a few human studies have been conducted to assess health effects associated with exposure to boron compounds, including boric acid.
Fatalities among young children have resulted from skin absorption of boric acid used as dusting powder on diapers; in 120 reported cases of poisoning, the mortality was 52.5% (Deichmann & Gerade 1969).
Accidental use of boric acid solution in the preparation of baby formula has resulted in poisoning in infants and a lethal dose of 2 to 3 g has been reported. Based on the reported lethal doses, which are not well documented in the literature, infants appear to be more sensitive than adults to boron compounds. (WHO 1998b).

After repeated oral administration to experimental animals, growth inhibition, organ weight changes, and testicular damage are the most striking effects observed.
The testis is the critical target organ with adverse effects being observed ranging from inhibited spermiation to degeneration of the seminiferous tubules with variable loss of germ cells, to complete absence of germ cells resulting in atrophy and transient or irreversible loss of fertility but not of mating behaviour.
Developmental toxicity has also been demonstrated in experimental animals with lower foetal bodyweight being the critical effect
(WHO 1998b).

5.14 Nitrate and nitrite

Nitrate and nitrite are naturally occurring ions that are part of the nitrogen cycle. The nitrate ion is the stable form for oxygenated systems although it can be reduced to nitrite by microbial action.
In general, vegetables will be the main source of nitrate intake in the general population. When the nitrate concentration in drinking water exceeds 50 mg/l (the limit value in Denmark), drinking water will be the major source of total nitrate intake, especially for bottle-fed infants. For the bottle-fed infant, daily intake from formula made with water containing 50 mg/l of nitrate would average about 8.5 mg/kg b.w. per day.
(WHO 1996).

Ingested nitrate is readily and completely absorbed from the gastrointestinal tract and rapidly distributed throughout the tissues. Approximately 25% of ingested nitrate is actively secreted into saliva, where it is reduced to nitrite by the oral microflora. Bacterial reduction of nitrate may also take place in other parts of the human gastrointestinal tract. (WHO 1996).

The toxicity of nitrate to humans is thought to be solely the consequence of its reduction to nitrite. The major biological effect of nitrite in humans is its involvement in the oxidation of normal haemoglobin to methaemoglobin, which is unable to transport oxygen to the tissues. The reduced oxygen transport becomes clinically manifest when methaemoglobin concentrations reach 10% of that of haemoglobin and above; the condition, called methaemoglobinaemia and also known as the "blue baby syndrome", causes cyanosis and, at higher concentrations, asphyxia. The haemoglobin of new-borns and young infants is more susceptible to methaemoglobin formation than that of older children and adults. (WHO 1996).

5.15 References

Andrews LS and Snyder R (1992). Ethyl alcohol. In: Casarett and Doull’s toxicology. The basic science of poisons. Fourth edition. Eds.: Amdur MO, Doull J and Klaassen CD. Pergamon Press. 412, 696-701.

ATSDR (1998). Toxicological Profile for chlorinated dibenzo-p-dioxins (Update). U.S. Department of Health & Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR (1997). Toxicological Profile for alpha-, beta-, gamma- and delta-hexachlorocyclohexane. Draft for Public Comment. U.S. Department of Health & Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

Braun-Fahrländer C, Künzli N, Domenighetti G, Carell CF and Ackermann-Liebrich U (1994). Acute effects of ambient ozone on respiratory function of Swiss schoolchildren after 10-minute heavy exercise. Pediatr Pulmonol 17, 169-177.

Burnett RT, Dales RE, Raizenne ME, Krewski D, Summers PW, Roberts GR, Raad-Young M, Dann T and Brook J (1994). Effects of low ambient levels of ozone and sulfates on the frequency of respiratory admissions to Ontario Hospitals. Environ Res 65, 172-194.

Cordier S, Garel M, Amiel-tison C, Morcel H and Mandereau L (1999). Neurologic and neurodevelopmental investigations of methylmercury-exposed children in French Guiana children. Neuroepidemiology 10, 102.

CSTEE (2000). CSTEE opinion on lead - Danish notification 98/595/DK. The Scientific Committee on Toxicity, Ecotoxicity and the Environment (CSTEE) of DG XXIV, Consumer Policy and Consumer Health Protection, Brussels, 5th May 2000.

CSTEE (1998). CSTEE opinion on phthalate migration from soft PVC toys and childcare articles. The Scientific Committee on Toxicity, Ecotoxicity and the Environment (CSTEE) of DG XXIV, Consumer Policy and Consumer Health Protection. November 1998.

Davidson PW, Myers GJ, Cox C, Axtell C, Shamlaye C, Sloane- Reeves J, Cernichiari E, Needham L, Choi A, Wang Y, Berlin M and Clarkson TW (1998). Effects of prenatal and postnatal methylmercury exposure from fish consumption on neurodevelopment. JAMA 280, 701-707.

Davis DL and Ahmed AK (1998). Exposures from indoor spraying of chlorpyrifos pose greater health risks to children than currently estimated. Environ Health Perspect 106, 299-301.

Darnerud PO, Eriksen GS, Jóhanesson T, Larsen PB and Viluksela M (1998). Polybrominated diphenyl ethers: Fodd contamination and potential risks. Nordic Council of Ministers. TemaNord 1998:503.

Deichmann WB and Gerade HW (1969). Toxicology of drugs and chemicals, Academic Press.

Eriksson P, Jakobson E and Fredriksson A (1998) Developmental neurotoxicity of brominated flame-retardants, polybrominated diphenyl ethers and tetrabromo-bis-phenol A. Proceedings from Polymer Additives and Monomers, Organohalogen Compounds 35, 375-377.

Gallo MA and Lawryk NJ (1991). Parathion, methyl parathion. In: Handbook of Pesticide Toxicology, vol. 2, Hayes WJ and Laws ER eds., Academic Press Inc., New York, 1040-1049, 985-987.

Goth A (1978). Chloramphenicol. In: Medical pharmacology, principles and concepts, 9th ed., The C.V. Mosby Company, Saint Louis.

Grandjean P and White RF (in press). Developmental effects of environmental neurotoxicants. In: Carroquino MJ, Hertz-Picciotto I, Bertollini R, eds. Children’s Health and the Environment. Rome: World Health Organization (in press).

Grandjean P, White RF, Nielsen A, Cleary D and de Oliveira Santos EC (1999). Mercury neurotoxicity in Amazonian children downstream from gold mining. Environ Health Perspect 107, 587-591.

Grandjean P, Weihe P, White RF, Debes F, Araki S, Murata K, et al. (1997). Cognitive deficit in 7-year-old children with prenatal exposure to methylmercury. Neurotoxicol Teratol 19, 417-428.

Gurunathan S, Robson M, Freeman N, Buckley B, Roy A, Meyer R, Bukowski J and Lioy PJ (1998). Accumulation of chlorpyrifos on residential surfaces and toys accessible to children. Environ Health Perspect 106, 9-16.

Harada M (1995). Minamata disease: methylmercury poisoning in Japan caused by environmental pollution. Crit Rev Toxicol 25, 1-24.

Hoek G, Brunekreef B, Kosterink P, van den Berg R and Hofschreuder P (1993a). Effect of ambient ozone on peak expiratory flow of exercising children in the Netherlands. Archiv Environ Health 48, 27-32.

Hoek G, Fischer P, Brunekreef B, Lebret E, Hofschreuder and Mennen MG (1993b). Acute effects of ambient ozone on pulmonary function of children in the Netherlands. Amer Rev Resp Diseases 147, 111-117.

Hunter DL, Lassiter TL and Padilla S (1999). Gestational exposure to chlorpyrifos: comparative distribution of trichloropyridinol in the fetus and dam. Toxicol Appl Pharmacol 158, 16-23.

IARC (1986). Tobacco smoking. IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Humans. Volume 38. IARC, Lyon, France.

Igata A (1993). Epidemiological and clinical features of Minamata disease. Environ Res 63, 157-169.

ILSI (1999). Overview of the health issues related to alcohol consumption. Executive summary of the book "Health issues related to alcohol consumption", 2nd edition.ILSI Europe Report Series.

Jedrychowski W, Flak E and Mróz E (1999). The adverse effect of low levels of ambient air pollutants on lung function growth in preadolescent children. Environ Health Perspect 107, 669-674.

Johnson BL, Hicks HE and De Rosa CT (1999). Key environmental human health issues in the Great lakes and St. Lawrence River Basins. Environ Res Section A 80, S2-S12.

Kinney PL, Thurston GD and Raizenne M (1996). The effects of ambient ozone on lung function in children: A reanalysis of six summer camp studies. Environ Health Perspect 104, 170-174.

Krzyzanowski M, Quackenboss JJ and Lebowitz MD (1992). Relation of peak expiratory flow rates and symptoms to ambient ozone. Arch Environ Health 47, 107-115.

Liu J, Olivier K and Pope CN (1999). Comparative neurochemical effects of repeated methyl parathion or chlorpyrifos exposures in neonatal and adult rats. Toxicol Appl Pharmacol 158, 186-196.

Loomis D, Castillejos M, Gold DR, McDonnell W and Borja-Aburto V (1999). Air pollution and infant mortality in Mexico City. Epidemiol 10, 118-123.

Lægemiddelkataloget (1996a). Chloramphenicol, 264. In Danish.

Lægemiddelkataloget (1996b). Sulfonamides, 276-277. In Danish.

Lægemiddelkataloget (1996c). Diethylstilboestrol, 569; oestrogens, 216-217. In Danish.

Marsh DO, Myers GJ, Clarkson TW, Amin-Zaki L, Tikriti S and Majeed M (1990). Fetal methylmercury poisoning: clinical and pathological features. Ann Neurol 7, 348-353.

McCance KL and Huether SE (1998). Pathophysiology. The biologic basis for disease in adults and children. Third edition. Mosby. 1214.

Murata K, Weihe P, Renzoni A, Debes F, Vasconcelos R, Zino F, Araki S, Jørgensen PJ, White RF and Grandjean P (1999). Delayed evoked potentials in Madeiran children exposed to methylmercury from seafood. Neurotoxicol Teratol 21, 343-348.

Mylchreest E and Foster PMD (1998). Antiandrogenic effects of di(n-butyl) phthalate on male reproductive development: a nonreceptor-mediated mechanism. CIIT 18(9).

Nielsen E (1997). Evaluation of health hazards by exposure to copper and estimation of a limit value in drinking water. Instituttet for Toksikologi, Levnedsmiddelstyrelsen. Baggrundsrapport udarbejdet for Miljøstyrelsen.

Nielsen E and Larsen PB (1996). Toxicological evaluation and limit values for DEHP and phthalates other than DEHP. Environmental Review No. 6 1996, Danish Environmental Protection Agency.

Norén K and Meironyté D (1998). Contamints in Swedish human milk. decreasing levels of organochlorine and increasing levels of organobromine compounds. Organohalogen compounds 38, 1-4.

Petersen JH (1999). Forurening af fødevarer med blødgørere - migration fra plast og generel baggrundsforurening. Ph.D.-afhandling, Afdelingen for Kemiske Forureninger, Instituttet for Fødevareundersøgelser og Ernæring, Fødevaredirektoratet.

Schardein JL (2000a). Alcohol use and abuse. In: Chemically induced birth defects 3rd ed., Marcel Dekker, New York; 735-743.

Schardein JL (2000b). Tobacco smoking. In: Chemically induced birth defects 3rd ed., Marcel Dekker, New York; 716-722.

Schardein JL (2000c). Antibiotics. In: Chemically induced birth defects 3rd ed., Marcel Dekker, New York; 383.

Schardein JL (2000d). Sulfonamides. In: Chemically induced birth defects 3rd ed., Marcel Dekker, New York; 380-382.

Schardein JL (2000e). Estrogens. In: Chemically induced birth defects 3rd ed., Marcel Dekker, New York; 292-297.

Stern BR, Raizenne ME, Burnett RT, Jones L, Kearney J and Franklin CA (1994). Air pollution and childhood respiratory health: Exposure to sulfate and ozone in 10 Canadian rural communities. Environ Res 66, 125-142.

Steuerwald U, Weihe P, Jørgensen PJ, Bjerve K, Brock J, Heinzow B, Budtz-Jørgensen E and Grandjean P (2000). Maternal seafood diet, methyl mercury exposure, and neonata neurological function. J Pediatr 136, 599-605.

WHO (1998a). Copper. Environmental Health Criteria 200. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1998b). Boron. Environmental Health Criteria 204. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1997a). Di-n-butyl phthalate. Environmental Health Criteria 189. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1997b). Flame retardants: a general introduction. Environmental Health Criteria 192. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1996). Lead. Nitrate and nitrite. In: Guidelines for drinking-water quality. Second edition, Vol. 2. World Health Organization, Geneva, 254-275 (lead), 313-324 (nitrate and nitrite).

WHO (1995). Inorganic lead. Environmental Health Criteria 165. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1994). Brominated diphenyl ethers. Environmental Health Criteria 162. World Health Organisation, International Programme on Chemical Safety, Geneva.

WHO (1993a). Polychlorinated biphenyls and terphenyls (second edition). Environmental Health Criteria 140. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1993b). Methyl parathion. Environmental Health Criteria 145. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1991a). Inorganic mercury. Environmental Health Criteria 118. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1991b). Lindane. Environmental Health Criteria 121. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1990). Methylmercury. Environmental Health Criteria 101. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1986a). Organophosphorus insecticides: a general introduction. Environmental Health Criteria 63. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1986b). Carbamate pesticides: a general introduction. Environmental Health Criteria 64. World Health Organization, International Programme on Chemical Safety. Geneva.

WHO (1984). Paraquat and diquat. Environmental Health Criteria 39. World Health Organization, International Programme on Chemical Safety. Geneva.

Woodruff TJ, Grillo J and Schoendorf KC (1997). The relationship between selected causes of postneonatal infant mortality and particulate air pollution in the United States. Environ Health Perspect 105, 608-612.