Perfluorinated Organic Compounds and Polybrominated Diphenyl Ethers Compounds – Levels and Toxicity in Aquatic Environments: A Review

Organochlorine pesticides are well known by the scientific literature to be persistent in the environment, toxic for wildlife and potential dangerous for humans since the publication of the famous volume "Silent Spring" by Rachel Carson in 1962 [1]. Due to their massive indus‐ trial production and commercialization for various human purposes, these chemicals have reached concentrations in worldwide environments that are able to significantly affect ter‐ restrial and marine wild species [2], remote world habitats [3] including the remote deep-sea [4], and protected species listed in the IUCN Red List of Threatened Species [5]. Actual con‐ centrations are able to severely affect trophic webs [6] and top-predators [7; 8]. As animal, humans are not excluded by the effects of pollution and concerning these compounds, feed‐ ing represents the principal and worldwide diffuse exposure mechanism for human popula‐ tions rather than inhalation and dermal contact [9; 10; 11; 12]. In spite of that, only in USA, chemical industry produces about 70,000 new products and organic chemicals accounted for the greatest share of production (364.2 million tons) in 1997 [13].


Introduction
Organochlorine pesticides are well known by the scientific literature to be persistent in the environment, toxic for wildlife and potential dangerous for humans since the publication of the famous volume "Silent Spring" by Rachel Carson in 1962 [1]. Due to their massive industrial production and commercialization for various human purposes, these chemicals have reached concentrations in worldwide environments that are able to significantly affect terrestrial and marine wild species [2], remote world habitats [3] including the remote deep-sea [4], and protected species listed in the IUCN Red List of Threatened Species [5]. Actual concentrations are able to severely affect trophic webs [6] and top-predators [7; 8]. As animal, humans are not excluded by the effects of pollution and concerning these compounds, feeding represents the principal and worldwide diffuse exposure mechanism for human populations rather than inhalation and dermal contact [9; 10; 11; 12]. In spite of that, only in USA, chemical industry produces about 70,000 new products and organic chemicals accounted for the greatest share of production (364.2 million tons) in 1997 [13].
The continuous research of new substances able to cover the great request from engineering, chemistry, pharmaceuticals, medical, commercial and social activities led to the direct and indirect release in environment and the consequent exposure of living organisms to new compounds. Once released in the environment, new chemicals interact with the abiotic and biotic matrices producing mixture composed by pure chemicals which auto-interact, their metabolites and/or reaction and degradation by-products. These mixture are characterized by a progressive increase of complexity and by a clear geographical footprint with percentages of chemical composition that are dependent both by physico-chemical properties of The following paragraphs aims to summarized actual knowledge on PFCs and PBDEs principal characteristics including environmental levels and toxicity on biota.

Physico-chemical properties of considered molecules
PFCs and PBDEs include molecules characterized by a similar chemical formula but also by very different physico-chemical properties. As consequence of the structural dissimilarities, differences concerning environmental distribution dynamics, and levels in abiotic and biological matrices are observed among PFCs and PBDEs congeners. Furthermore, the ecotoxicological risk associated to the diffusion of these persistent organic pollutants could be notably dissimilar. In fact, physico-chemical properties of molecules influence possible adverse effects on non-target biological communities. In addition, observed toxicity is notably affected by the interaction among considered chemicals and environmental matrices caused by the photo-chemical deterioration and the production of metabolites during microbial biodegradation phenomena.

Perfluorinated organic compounds (PFCs)
Concerning chemicals of ecotoxicological interest, perfluorinated organic compounds (PFCs) are of particular emerging interest due their documented presence both in wildlife's tissues and human blood PFCs [25].
PFCs are anionic, and fluorine-containing surfactants (both soluble in water and oil) and are applied for a large industrial and commercial employment to produce surfactants, lubricants, paints, polishes, food packaging, and fire-fighting, foams propellants, agrochemicals, adhesives, refrigerants, fire retardants, and medicines [26; 27].
Their structure consisting of a fluorine atom with which all hydrogen atoms from the linearalkyl chain, which is a hydrophobic group, are replaced. Physico-chemical properties of PFCs favour the occurrence of long-range transport dynamics, as they are more volatile than chlorine or bromine analogues.
Among PFCs, perfluorooctanoic acid (PFOA) and perfluoroctanesulfonic acid (PFOS) represents the principal compounds of environmental concern.
Salts of perfluorooctanoic acid (PFOA, C 8 HF 15 O 2 ) have been used as surfactants and processing aids in the production of fluoropolymers, and these salts are considered critical to the production of certain fluoropolymers and fluoroelastomers [28]. The functional chemical structure is C 7 F 15 COOH and for this reason tends to behavior like an acid dissociating as follows: C 7 F 15 COO -+ H + . Perfluorooctane sulfonate (PFOS; C 8 HF 17 O 3 S even in this case it dissociates as follows: C 7 F 17 SO 3 -+ H + ) evidences an excellent chemical and thermal stability and is a chemical precursor for the synthesis of other molecules [26] such as fluorinated surfactants and pesticides (Abe and Nagase., 1982 in [29]). Perfluoroalkanesulfonate salts and perfluorocarboxylates are reported to be present in fire-fighting foam formulations, including aqueous film forming foams which are fire-fighting materials largely used by military bases and airports to face hydrocarbon fuel fires or to prevent the potential risk of fire [30; 31]. Moody and colleagues [32] reported for the 2001, an estimated PFOS annual production quantity in United States of America of 2,943,769 kg.
Principal PFOA and PFOS chemical properties are summarized in Table 1.

Polybrominated Diphenyl Ethers (PBDEs)
Polybrominated diphenyl ethers (PBDEs) are a class of organohalogen compounds used worldwide over the past three decades as chemical additives to reduce the flammability of common use products [34]. These chemicals were first introduced to the market in the 1960s and their global demand has increased rapidly since the end of the 1970s, due to the growing popularity of personal computers and other electronic equipment, to which they were added to improve fire safety [35].
Since '70 PBDEs were used as flame retardants in a wide range of common use such as cloths, foam cushions, polyurethane sponges, carpet pads, chairs, couches, electronic instruments including computer castings, and insulating materials [36].
In 2000, the industrial production of these chemicals has been esteemed to be around the 64,000 cubic tons per year. The 50% of this annual production was commercialized in America, while in Europe only 12% [37].
Because of toxicity and persistence of PBDEs, these chemicals are included in the persistent organic pollutants (POPs) Reviewing Committee (www.pops.int) and their industrial production is to be eliminated under the Stockholm Convention.
Polybrominated diphenyl ethers are, apart from the oxygen atom between the phenyl rings, structurally similar to PCBs, consisting of two halogenated aromatic rings linked by an ether group. PBDEs chemical synthesis is performed by the diphenyl-ethers bromination in presence of dibromomethane as solvent. Diphenyl-ethers have 10 hydrogen atoms and each of them can be replaced by an atom of bromine. This reaction could produce 209 possible congeners, numbered from 1 to 209 in relation to the number of bromine atoms substituting hydrogen ones and their relative position within the molecule [38]. The available commercial PBDE products are not single compounds or even single congeners but rather a mixture of congeners. Nevertheless, commercial mixtures are constituted by a little part of the 209 possible congeners due to the instability of a large part of them [39] which tend to quickly debrominate.
PBDEs are semi volatile compounds characterized by a low vapor pressure and a scarce water solubility. These properties tends to decrease with the level of substitutions by bromine atoms in the molecular structure whereas hydrophobic properties increase. Octanol/water distribution coefficients (K ow ) are variable with substitutions and are included within: 5.9-6.2 for tetra-BDE, 6.5-7.0 for penta-BDE, 8.4-8.9 for octa-BDE, and 10.0 for deca-BDE. PBDEs half-life in air are extimated to be about two days, while in water longer times are modeled (two months) whereas in soils and sediments average half-lives are six months [40].

Sources, distribution dynamics, and environmental levels
Concerning PFCs, principal environmental sources are represented by the direct diffusion of surfactants, lubricants, paints, polishes, foams propellants, agrochemicals, adhesives, refrigerants, fire retardants, and medicines containing these chemicals. Indirect releases could occurs from food packaging and painted manufacturing when discharged and exposed to rain and bad weather conditions. Nevertheless, the large use of fire-fighting materials containing PFCs both when a critical fire occurs and to prevent accidents in high risk procedure (i.e. military or firemen exercitations, routine activities, airports activities), represents the principal direct diffusion of these chemicals on the ground able to affect wide geographical surfaces, superficial and groundwater [31].
As regard as PBDEs, environmental releases could occurs during manufacturing lifetimes. Releasing mechanism are not completely cleared, however, it is believed that PBDEs are released to the air when objects are manufactured and during object's life span. Their disposal and waste could produce releases too [42]. In the last years recycling of end products containing PBDE is becoming the principal source of release of these chemicals in the environment [43]. Burning of plastics, waste electronic goods, and oil shale may provide an additional PBDEs loads both in atmosphere and soil. Also, productive processes represents an important source, high levels are measured in environmental matrices closed to the flame-retardants factories [40].
Monitoring PFCs and PBDEs in environmental matrices evidenced first of all the needing to develop accurate sampling strategies to collect representative samples from heterogeneous and quickly variable matrices such as air and water are. On the contrary, soils and sediments even if much more stable present structural heterogeneity (i.e. organic matter content and composition, grain-size structure, redox conditions) which could interfere with quantifications and data interpretation. Concerning biota the matter (if it is possible!) is quite more complex. Measured levels could be affected by a lot of different factors as well as age, sex, phase of animal life-stage, lipid content, water content in tissues, part of the animal excised for the analyses and much more other factors. Another point is represented by the sampling treatments and the detecting method adopted to perform laboratory analyses. Different methods are associated to different detection limits, precision and accuracy. Low polluted matrices such as air and water required methods able to detect levels of chemicals at concentrations measured in pg/L, whereas biological tissues allowed the adoption of quite less sensible methods as well as concentrations are usually measured in mg/kg or ng/g.
Hereby levels reported by the literature in different environmental matrices are reported organizing them per matrix. When possible information about the sampling strategies adopted are reported (i.e. depth of sampling for water and soils or sediments, geographical areas, type of tissues), nevertheless a complete data selection related to the sampling strategies, sampling treatments, and detecting methods has not been possible due to the wide heterogeneity in data acquisition procedures.

Air
Low data are available on air levels, probably due to the great difficult associated to the sampling of this matrix and samples treatment strategies in laboratory. Laboratory (i.e. air, laboratory rooms, instruments, vials, etc.) and cross-over contaminations are extremely simple to occur treating air samples. Furthermore, adopted methods have to be extremely sensitive.
In the period from 1994 to 1995, measured levels of total PBDEs (congeners not specified) reached maximum values of 28 pg/m 3 in samples collected in Alert, Canadian Arctic [36]. The rural area of southern Ontario showed in the Early spring of 2000, notably higher levels of total PBDEs (as sum of 21 congeners detailed in the paper) ranging within 10-1,300 pg/m 3 [45].

Terrestrial environments and soils
Soil pollution could derived by direct local sources but also by dry-air depositions or runoffs. Humus represents the soil fraction able to accumulate chemicals due to the presence of both hydrophobic and hydrophilic molecules. From here chemicals could be re-volatilized in air, transferred throughout the soil trophic web or be leached throughout rains affecting groundwater ecosystems with possible important consequences for humans. The net dominance of one of these phenomena is a factor dependent to the geographical position of the area (affecting air/soil temperature, sun irradiance, quantities of rains, etc.) and to the soil physico-chemical characteristics.
In soils sampled closed to a polyurethane foam manufacturing facility in the United States, concentrations of total PBDEs (tetra-and penta-BDE) of 76 µg/kg dry weight are reported. Average values measured in soil downwind from the facility were significantly lower 13.6 µg/kg dry weight [48; 49].
In a study performed throughout the Estonian State, PBDEs levels are defined in soils for the first time. Total values observed ranging within <0.01-3.2 ng/g (d.w.) as reported by Kumar and colleagues [50]. Even if measured values are not excessive, authors predict a possible increase in the near future due to the particular waste policy of the Estonia.
In Australia, superficial samples collected in 39 remote, agricultural, urban and industrial locations from all states and territories evidenced highest (but not indicated) values from "urban and industrial areas, particularly downstream from sewage treatment plants" (Australian Government, on-line available at: http://www.environment.gov.au/settlements/publications/ chemicals/bfr/pubs/factsheet.pdf.). In the same report, levels are indicated to be comparable to values measured in other European and Asian Countries.

Aquatic environments
Aquatic ecosystems represent the final reservoir for PFCs and PBDEs due to their great affinity towards sedimentary and living organic matter. In these systems, measured levels of POPs could increase along the trophic web affecting humans feeding aquatic species.
A recent study performed by Nakata and colleagues [51] evidenced significant differences among tidal and coastal levels of PFOA and PFOS in all considered environmental matrices supporting the existence of different dynamics affecting PFCs distribution and ingress in trophic webs that are zone dependant in marine ecosystems.
Even if some researchers, as reported below, have been performed to evaluate PFCs and PBDEs levels in environmental matrices and biota from river and marine ecosystems, no data are available on environmental levels, bioaccumulation and biomagnification dynamics occurring in coastal lagoons and transitional areas which are completely non explored. This lacking in scientific data could affect risk evaluations linked to human exposure to these chemicals in transitional areas. In fact lagoons and estuaries are the most populated, polluted and productive areas in the world. Feeding exploitation of these not explored ecosystems could represent a notable and not considered risk for humans.

Sediment
In Japan aquatic environments, Senthilkumar and colleagues [44] observed PFOA measurable levels only in sediments sampled from the Kyoto river ranging within 1.3-3.9 ng/g dry weight (dw) and not measurable PFOS levels.
Becker and colleagues [59] evidenced that once released in water, PFCs accumulate into sediments with a PFOA/PFOS ratio of about 10. In particular, PFOA were 10-fold less than PFOS but enrichment observed on sediment was not correlated to the total organic carbon contents.
From several sites sampled along the Columbia River system, in south eastern British Columbia, Rayne and colleagues [61] measured PBDE concentrations (as sum from di-to penta-BDE congeners) included within 2.7-91 µg/kg.
Sediments from two Arctic lakes in Nunavut Territory evidenced measurable concentrations from 0.075 to 0.042 µg BDE 209/kg dw. One of the two Arctic lakes sampled was located near an airport and PBDEs inputs from this source could not be excluded [62]. Authors hypothesized a particles-mediated transport to the Canadian Arctic due to its low vapour pressure and high octanol-water partition coefficient.

Biota
POPs could accumulate in species evidencing interspecies differences as well as sex and size-related ones [64]. Recent studies evidenced that POPs concentrations in demersal fishes varies significantly relating to the sex, maturity, and reproduction [65].
Data collected in fishes and fishery products [52]  In the same research, PFOS levels measured in liver samples collected from ringed and gray seals (Bothnian Bay, Baltic Sea) range within 130-1,100 ng/g, w.w.. In this case, no relationships are observed between PFOS levels and ringed or gray seals age but levels measured in livers are 2.7-5.5 fold higher than values in blood with a positive strong correlation between blood and liver levels. Concerning white-tailed sea eagles (Germany and Poland) indicate increasing of concentrations from 1979 to 1990s. Livers of Atlantic salmons do not evidenced measurable levels neither PFOS nor PFOA.
In 2007, Senthilkumar and colleagues [44] define levels of PFOA and PFOS in biotic compartment of aquatic ecosystems in Japan. Concerning fish tissues, only jack mackerel showed PFOA and PFOS respectively at averages of 10 and 1.6 ng/g w.w.. Wildlife livers contained PFOS levels ranging within 0.15-238 ng/g w.w. and PFOA values included within <0.07-7.3 ng/g w.w.. Cormorants showed maximum accumulation followed by eagle, raccoon dog and large-billed crow.
Kannan and colleagues [73] measured PFOA and PFOS levels in livers of birds collected from Japan and Korea (n= 83). PFOS was found in the livers of 95% of the birds analyzed at concentrations greater than the limit of quantitation (LOQ=10 ng/g, w.w.). The greatest concentration of PFOS of 650 ng/g, w.w., was found in the liver of a common cormorant from the Sagami River in Kanagawa Prefecture.
Borghesi and colleagues [24] evidenced a PBDEs concentrations in Antarctic fish species ranging within average of 0.09 ng/g (w.w.) recorded in G. nicholsi to average of 0.44 ng/g (w.w.) measured in C. gunnari. In Mediterranean tuna PBDEs levels were two or three orders of magnitude higher (15 ng/g w.w.). Furthermore, PBDE congener profiles differ between species; low brominated congeners prevailed in Antarctic species while in tuna tetraand penta-bromodiphenyl ethers are the most abundant groups (41% and 44%, respectively). In the same study, a strong correlation with the fish length is observed for the species C. hamatus but the same relation is not recorded considering the weight. Tuna evidences a gender dependency in PBDEs concentrations in fact levels are significantly high in females than in males (18 ng/g vs 13 ng/g w.w.) which authors attribute to the lower fat content in males.

Human exposure
PBDEs levels in humans have increased over the past several decades. Schiavone and colleagues [74] measured PBDEs in human lipid tissues from Italy even if at values lower than other POPs (PCBs and DDTs). These chemicals are structurally similar to thyroid hormones (i.e. thyroxine T4) and could disrupt thyroid homeostasis as observed in laboratory experiments on animals [75] causing damages similar to thyroid hormone deficiencies [76; 77].
Effects on male reproductive system have been documented by the literature due to the weak estrogenic/antiestrogenic activity of these chemicals [78]. In rats, the exposure to a single dose of 60 µg/kg body weight (b.w.) of PeBDE-99 produces significant decreases of sperm numbers. Akutsu and colleagues [79] evidenced relationship between human serum PBDEs and sperm quality. In particular, PBDE levels in Japan men are comparable to those found in European countries and a strong inverse correlations were ob-served between the serum concentration PeBDE-99 and sperm concentration (r = -0.841, p = 0.002) and testis size (r = -0.764, p = 0.01).
The Department of the Environment and Water Resources of Australia founded in 2004 a research aimed to evaluate PBDEs levels in indoor environments collecting and analysing samples from air, dust and surfaces from homes and offices in south-east Queensland. Concentrations of PBDEs were greater in indoor air than in outdoor once, evidencing that major risks are related to the indoor exposure. Furthermore, the lowest PBDE concentration in indoor dust was found in a house with no carpet, no air-conditioning, and which was older than five years. The highest concentration was found in an office with carpet and air-conditioning, and which had been refurbished in the last two years. A recent study developed by Meeker et al. [77] evidenced altered serum hormone levels in US men affected by infertility clinic (n=24) as a result of indoor exposure to PBDE. BDE 47 and 99 were detected in 100% of dust samples, and BDE 100 was detected in 67% of dust samples. A significant inverse relationship between dust PBDE concentrations and free androgen index was observed. Furthermore, dust PBDE concentrations were inversely associated with luteinizing hormone (LH) and follicle stimulating hormone (FSH), and positively associated with inhibin B and sex hormone binding globulin (SHBG).
Concerning POPs levels in humans, infants show the higher feeding exposure compared to adults due to their high feed consumption per kilogram of body weight. Weijs et al., [80] evidenced in not-breastfed Dutch infants a progressively increasing exposure to POPs during the first year growth from the birth due to the diet changes. Concerning PBDEs, the mean level measured in breast milk was 3.93±1.74 ng/g lipid and the estimated PBDE daily intake for a breastfed infant was 20.6 ng/kg b.w./day after delivery [81].
Chao and colleagues [82] evidenced that, in Taiwan, PBDEs levels in breast milk (n= 46) are associated with demographic parameters, socioeconomic status, lifestyle factors, and occupational exposure. Average levels measured in 2010 (1.07-3.59 ng/g lipid) were 0.7-fold lower than in 2000. Furthermore higher levels of PBDEs were positively correlated to the maternal age and are not correlated with maternal pre-pregnant BMI (Body mass index), parity, and lipid contents of breast milk.
PBDEs level in breast milk is lower in more educated women after controlling for age and pre-pregnancy BMI in tested mothers, nevertheless these results are not completely in agreement with Wang and colleagues [83] which evidenced that the mean level of BDE47 in breast milk from mothers with pre-pregnant BMI <22.0 kg/m 2 had a significantly higher magnitude compared to those with pre-pregnant BMI >22.0 kg/m 2 (1.59 vs 0.995 ng/g lipid, p= 0.041) and no relationships between PBDEs exposure levels and women's age, parity, blood pressure, annual household income, and education level.
Evidences regarding a relationship between PBDEs levels in breast milk and seafood consumptions in Taiwan has been explored by the literature [83]. Women eating more fish and meat show not significantly higher PBDE levels than others nevertheless, a significant difference in PBDE levels was demonstrated between the higher (2.15 ng/g lipid) and lower (3.98 ng/g lipid) shellfish consuming subjects (p = 0.002) after an adjustment for the confounders.

Ecotoxicological effects
Data collected by 3M [84; 85] on ecotoxicological effects of PFOS on aquatic species evidenced after 96h of exposure an E b C 50(biomass) of 71 mg/L, an E g C 50(growth rate) of 126 mg/L, and a NOEC (biomass/growth rate) of 48 mg/L on Selenastrum capricornutum (Algae). Acute effects on the oyster shell deposition are observed at 2.1 mg/L (96h NOEC) whereas subchronic/chronic effects were recorded on Mysidopsis bahia (mysid. shrimp) at 0.25 mg/L (35-day NOEC reproduction/growth). The same species evidenced respectively acute effects at 4.0 mg/L (96h EC 50 ) and 1.2 mg/L (96h NOEC).
The species Crassostrea virginica evidenced acute toxicity at levels higher than 3.0 mg/L (96h EC 50 ) and 96h NOEC at 1.9 mg/L.
Acute toxicity is observed in Daphnia magna exposing animals at 66 mg/L (48h EC 50 ) whereas chronic effects are observed over 7 mg/L (28day NOEC reproduction ).
Sanderson and colleagues [86] evidenced the ecological effects induced by the exposure to perfluorinated surfactants (PFOS and PFOA), on zooplankton species performing using classical ecotoxicological tests, 30-L indoor microcosm and 12,000-L outdoor microcosm experiments. The zooplankton community considered in this experiment was composed by the following representative species: Cyclops diaptomus, C. strenuus, Canthocamptus staphylinus, D. magna, Keratella quadrata, Phyllopoda sp., Echninorhynchus sp., Ostracoda sp., and total Rotifera sp. In addition to zooplankton and pond snails, occasional macrophytes (Elodea canadansis and Myriophyllum spicatum), and larger invertebrates (Ephemeroptera sp., Assellus aquaticus) were present.
Results evidenced that zooplankton had lower tolerance toward PFOS than toward PFOA. Researchers observed that "with increasing concentrations the zooplankton community became simplified toward more robust rotifer species, which, as an indirect effect, increased their abundance due to a shift in competition and predation". Functional studies evidenced that PFOS inhibits gap junction intercellular communication (GJIC) in rat liver epithelial cells cultured in vitro (personal comunication reported in [29]) and that it is an uncoupler of phosphorylation in rat liver mitochondria (personal comunication reported in [29]).
PBDEs can inhibit growth in colonies of plankton and algae and depress the reproduction of zooplankton.
Laboratory mice and rats have also shown liver function disturbances and damage to developing nervous systems as a result of exposure to PBDEs (http://www.environment.gov.au/ settlements/chemicals/bfrs/index.html).
Ecotoxicological tests performed on PBDEs on different species and medium following exposed expressing results as: LC 50(median lethal dose) , LOAEL (Lowest-Observed-Adverse-Effect Level) , LOEC (Lowest-Observed-Effect Concentration) , NOAEL (No-Observed-Adverse-Effect Level) , and NOEC (No-Observed-Effect Concentration) . The use of the letter "a" following data means that in the study reported highest concentration (or dose) tested did not result in statistically significant results. Since the NOEC or NOAEL could be higher, the NOEC or NOAEL are described as being greater than or equal to the highest concentration (or dose) tested.
The Great Lakes Chemical Corporation [98], reported ecotoxicological results obtained on the Rat. The exposure mixture PeBDE (DE-71) was composed by 45-58.1% of penta-BDE, 24.6-35% tetra-BDE [99; 100]. Exposure doses of PeBDE were 0, 2, 10 and 100 mg/kg bw per day (doses adjusted weekly based on mean body weight of animals) and after 90 days observed effects were: • decreased food consumption and body weight, increased cholesterol, increased liver and urine porphyrins at 100 mg/kg bw dose, • increased absolute and relative liver weights at 10 and 100 mg/kg bw, with return to normal ranges after 24-week recovery period, • compound-related microscopic changes to thyroid and liver at all dosage levels, • microscopic thyroid changes reversible after 24 weeks, • microscopic liver changes still evident at all dosage levels after 24-week recovery period, • liver cell degeneration and necrosis evident in females at all dosage levels after 24-week recovery, • LOAEL (liver cell damage) = 2 mg/kg bw, • NOAEL could not be determined, as a significant effect was observed at the lowest dose tested.
Little data have been gathered on the associations between PBDEs exposure and birth outcome and female menstruation characteristics in both epidemiological and animal studies.
In rats, the in utero exposure to PBDEs reduces the number of ovarian follicles in rat females and causes permanent effects on rat males [81].

Conclusions
Recent data collected on these chemicals evidence significant levels in environments, wildlife and humans. In particular, observed ecotoxicological effects on species, measured values in humans tissues and their relationships with fertility suggest that PFCs and PBDEs represent an important problem to be quickly solved. Unfortunately, collected data both on chemical distribution in abiotic and biotic matrices are fragmentary and incomplete as well as ecotoxicological studies on laboratory species and microcosms. Studies in aquatic ecosystems and, in particular, in transitional ones have to be improved to allow a correct evaluation of the exposure risk for humans to these compounds due to the dietary intakes.

Author details
Monia Renzi *