Perspective on halogenated organic compounds

Abstract

During the past century, a vast number of organic chemicals have been manufactured and used in industrial, agricultural, public health, consumer products, and other applications. The widespread use in bulk quantities of halogenated organic chemicals (HOCs; also called Organohalogens), including chlorinated, brominated, and fluorinated compounds, and their persistent nature have resulted in global environmental contamination. Increasing levels of HOCs in environmental media (i.e., air, water, soil, sediment) and in human tissues including adipose tissue, breast milk, and placenta continue to be a cause of ecological and human health concern. Human exposure can occur through multiple pathways including direct skin contact, inhalation, drinking water, and mainly through food consumption. HOCs exposure has been implicated in a myriad of health effects including reproductive, neurological, immunological, endocrine, behavioral, and carcinogenic effects in both wildlife and humans. In addition, recent studies indicate that exposure to HOCs contributes to obesity and type 2 diabetes. Because of these adverse health effects, several regulatory agencies either banned or placed severe restrictions on their production and usage. In turn, many industries withdrew from production and usage of HOCs. This action resulted in decline of older HOCs such as polychlorinated biphenyls (PCBs), but more recent HOCs such as polybrominated diphenyl ethers (PBDEs) and perfluoroalkyl substances (PFAS) show a steady increase/stable with time in the global environment. Based on their use pattern and their persistent chemical properties, human exposure to HOCs will likely continue. Hence, understanding human health effects and taking preventive measures for such exposures are necessary.

1. Introduction

Human activities alter the quality of the environment on a global scale, which can adversely affect life on earth. A prototypical example of such environmental damage and harmful biological effects is that caused by persistent human-made chemicals either intentionally or unintended byproducts, particularly halogenated organic compounds (HOCs). HOCs are organic compounds that contain one and up to several chlorines (named chlorinated, Fig. 1), bromines (named brominated, Fig. 2), and fluorines (named fluorinated, Fig. 3) with different positions leading to several congeners (Kodavanti and Loganathan, 2019; Kodavanti et al., 2022a, b). They are used widely in industries because of they are effective and inexpensive to manufacture.

Fig. 1.

Chemical structures of chlorinated HOCs. PCBs, polychlorinated biphenyls; Chlordane, DDT, 1,1,1-trichloro-2,2-bis[p-chlorophenyl]-ethane; HCH, hexa-chlorocyclohexane; PCDDs, polychlorinated dibenzodioxins; PCDFs, polychlorinated dibenzofurans, heptachlor, and toxaphene. The phenyl rings may have a variable number of chlorine atoms, from 1 to 10, in 209 possible combinations. The letters (o), (m), and (p) indicate ortho, meta, and para substitutions for chlorines and numbers indicate position of chlorine in the generalized structures.

Fig. 2.

Chemical structures of brominated HOCs. Core structure as well as some predominant PBDE congeners such as PBDE-47, PBDE-99, and PBDE 153. The phenyl rings may have a variable number of bromine atoms, from 1 to 10, in 209 possible combinations. The letters (o), (m), and (p) indicate ortho, meta, and para substitutions for bromines and numbers indicate position of bromine in the generalized structures.

Fig. 3.

Chemical structures of fluorinated HOCs. PFOS (perfluorooctane sulfonate), PFOA (perfluorooctanoic acid), Perfluorooctyl sulfonamides, where R = CH₂CH₃, CH₂CH₂OH, CH₂OH or H. An example of FTOH, heptadecafluoro-1-decanol (8:2 FTOH), HFPO-DA, hexafluoropropylene oxide dimer acid; GenX (HFPO-DA ammonium salt) are also shown here.

HOCs remain in the environment for extended periods of time, predicted to take decades or even centuries degrade. Several factors contribute to their persistence in the ecosystem. For example, chemicals are often degraded by ultraviolet light (UV) from the sun or oxidized in the atmosphere. HOCs resist degradation through these natural processes and can become concentrated in sediment, water or air. These compounds can be volatile (i.e., can vaporize in the air) or travel by water currents through the process of evaporation and re-deposition. These traits allow HOCs to be transported over long distances, far from the original source of contamination (Fig. 4). HOCs are removed from the atmosphere by physical processes such as wet dry deposition or vapor uptake and are then deposited on soils, surface waters and plant surfaces. Most of the HOCs are deposited onto surface waters absorb into suspended sediments. Once bound to soil and sediment, these chemicals generally remain fixed except for bulk transport due to soil erosion, flooding, and dredging (Dickson and Buzik, 1993).

Fig. 4.

Schematic showing HOCs in different parts of the environment and main environmental sources during long-range atmospheric transport, bioaccumulation, and biomagnification. Adapted from Kodavanti et al. (2014). In: Reference Module in Biomedical Research, 3rd ed, pp. 1–9, Elsevier. http://dx.doi.org/10.1016/B978-0-12-801238-3.00211-7.

Because of the lipophilic nature of HOCs, ingestion of these compounds builds up in fatty tissue of living organisms, bioaccumulates in the food chain (Fig. 4), and leads to serious health consequences for humans and wildlife. HOCs are detected in almost all tissue or environmental samples from almost every region of the world. This affinity for lipid rich tissue allows the HOCs to accumulate, persist due to their resistance to biological degradation, and bioconcentrate in the body. Consequently, even though the level of HOCs exposure may be limited, they can eventually reach toxicologically relevant concentrations due to continuous exposure.

Due to their extreme persistence in the environment, bioaccumulative nature, and long-term health effects in humans, some HOCs such as polychlorinated biphenyls (PCBs, an industrially versatile compound), insecticides such as DDTs, HCHs, chlordane, heptachlor, toxaphene, and industrial byproducts such as chlorinated/brominated dioxins/dibenzofurans are well-known global environmental contaminants (Loganathan et al., 2020; Kodavanti et al., 2022b). Although the use of some of these HOCs has been banned or severely restricted in most developed countries more than five decades ago, these HOCs are still found in almost all compartments of the global ecosystem and pose a threat to life. Following the ban on their production and use, residue levels of those HOCs have declined, but in some cases at a relatively slow rate. Nevertheless, new HOCs continue to be discovered in the environment. Per-and polyfluoroalkyl substances (PFAS), polybrominated diphenyl ethers (PBDEs), triclosan, triclocarban, tetrabromobisphenol A (TBBPA) and hexabromocyclododecanes (HBCDs) are widely used in a variety of industrial and consumer products and are considered as chemicals of environmental and health concern on the global scale. (Fig. 2, Fig. 3 and Fig. 5). This chapter contains basic information about HOCs explaining different classes, properties, widespread environmental contamination, and human/animal health effects focusing on chlorinated, brominated, and fluorinated compounds. Other chapters in this volume contain detailed information about these different classes of HOCs focusing on neurotoxic effects including neurobehavior, neurophysiological, neurohormonal, neurochemical, role of microbiome, and structure-activity relationship among these compounds.

Fig. 5.

Chemical structures of other emerging HOCs. Tetrabromobisphenol A (TBBPA), hexabromocyclododecane (HBCD), decabromodiphenyl ethane (DBDPE), hexabromobenzene are brominated HOCs while F53B and ADONA are fluorinated HOCs. Triclosan and triclocarban belong to chlorinated HOCs.

2. Background and classes of HOCs

Chlorinated HOCs were synthesized as early as 1830. Polychlorinated biphenyls (PCBs) were first synthesized in the early 1880s (Schmidt and Schultz, 1881), and their commercial production began in 1929. Commercial PCB formulations were sold under different trade names in the world. For example, in the United States and Great Britain the most common trade name for PCBs was Aroclor. PCB mixtures were named according to their chlorine content. For instance, Aroclor 1254 contains 54% chlorine by weight, and Aroclor 1260 contains 60%. The PCB mixture formulations were different depending on the country of origin and were produced in Germany (Clofen), Italy (Fenclor), France (Phenoclor and Pyralene), Japan (Kanechlor), Russia (Sovol), and Czechoslovakia (Delor). PCB mixtures were produced for a variety of uses such as fluids in electrical transformers, capacitors, heat transfer fluids, hydraulic fluids, lubricating and cutting oils, and as additives in plastics, paints, copying paper, printing inks, adhesives, and sealants. Other widely used HOC is the agricultural insecticide, 1,1,1-tri-chloro-2,2-bis[p-chlorophenyl]-ethane (DDT). It was first synthesized before the turn of the 20th century by Zeilder in 1874. Its insecticidal value was discovered in 1939 (by Paul Müller), and it was put to field use in the 1940s. Application of DDT contributed to rapid reduction of malaria and other insect-borne diseases such as typhoid fever and cholera. Such an advantage led to an astonishing increase in agricultural productivity in many regions of the world. The great success in the application of DDT during and following the Second World War in 1948 earned Paul Müller the Nobel Prize in Medicine. Other HOCs such as hexachlorocyclohexanes (HCHs) and chlordane were introduced in 1945 and these and other similar insecticides contributed to human welfare as agricultural and domestic pest control agents. Ultimately, however, owing to their persistent, bioaccumulative and toxic properties in the environment and the long-term human health effects, these chlorinated HOCs were banned or severely restricted in the early 1970s in several developed countries (Loganathan, 2012, 2016). Following this initial restrictive action, HOCs were the subject of a concerted regional, national, and international effort to limit their production and usage, and to control the disposal of materials no longer in use. Highly toxic HOCs such as polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs) have never been intentionally produced but are released into the environment as byproducts of combustion of chlorinated HOCs (Kodavanti et al. 2022b). PCDDs are composed of two benzene rings connected by two oxygen atoms and contain four to eight chlorines, for a total of 75 congeners (Fig. 1). PCDFs are also composed of two benzene rings. The rings have one oxygen molecule between them and have four chlorine binding sites available on each ring, making a total of 135 different PCDF congeners (Fig. 1) (Huwe, 2002; Kodavanti et al., 2014).

The next class of HOCs of concern is brominated compounds, including polybrominated biphenyls (PBBs), polybrominated diphenyl ethers (PBDEs), polybrominated dibenzo-p-dioxins/furans (PBDDs/PBDFs), tetrabromobisphenol A (TBBPA), hexabromobenzene, and hexabromocyclododecane (HBCD) (Fig. 2). PBBs and PBDEs constitute an important group of flame retardants (BFRs) (Kodavanti and Loganathan, 2019). It is estimated that approximately 13 million pounds of PBBs were produced in the USA from 1970 to 1976 and used for incorporation into plastic products that included business machine housings, radios, televisions, thermostats, electric shavers, hand tools and various automotive parts (Damstra et al., 1982; DiCarlo et al., 1978; Headrick et al., 1999). Three commercial PBB products were manufactured in the USA: hexabromobiphenyl, octabromobiphenyl and decabromobiphenyl (DiCarlo et al., 1978; Hardy, 2000). Hexabromobiphenyl was the predominant product with approximately 11.8 million pounds of chemicals being produced. Over 98% of the hexabromobiphenyl was produced as FireMaster BP-6 with the remainder being produced as FireMaster FF-1 (Hesse and Powers, 1978) after addition of an anti-caking agent to FireMaster BP-6. Production of PBBs ceased in 1974 (DiCarlo et al., 1978; Kodavanti and Loganathan, 2016).

PBDEs are added to consumer products to prevent combustion if exposed to flame or heat. PBDEs are added to plastics, upholstery, fabrics, and foam rubber and are found in common products such as computers, television sets, mobile phones, furniture, and carpet pads. Nearly 90% of electrical and electronic appliances contain PBDEs, which are added as flame retardants, affording up to 15 times greater escape time in case of a fire. PBDDs/PBDFs are relatively less toxic than chlorinated dioxins and are formed during heating or incineration of PBBs and PBDEs. Low levels of PBDDs/PBDFs detected in environmental samples suggest relatively lower exposure to biota (fish) and humans to these compounds.

In contrast to PCBs, some PBDEs are currently being produced and used in household materials. PBDEs are primarily indoor pollutants. PBDEs leach into the environment when household wastes decompose in landfills or are incompletely incinerated. Human health concerns stem from the fact that PBDEs (Fig. 2) are persistent, bioaccumulative, and structurally similar to PCBs (Fig. 1). PBDE concentrations are rapidly increasing in the global environment and in human blood, breast milk, liver, etc. Due to their ubiquitous nature in the environment and tendency to bioaccumulate in wildlife and humans, their exposure resulted in several toxic effects including the reproductive, hormonal, developmental, and nervous systems (Kodavanti and Curras-Collazo, 2010; Kodavanti and Loganathan, 2016).

Per- and polyfluoroalkyl substances (PFAS) nicknamed “forever chemicals” are another group of highly persistent HOCs (Jansen, 2019). PFAS are used in specialized consumer and industrial products. PFAS are used in metal-plating baths, surfactants, cleaning products, rust inhibitors, fire-fighting applications, starting materials for polymers, herbicide and insecticide formulations, cosmetics, shampoos, pharmaceuticals, water and oil repellent coatings for fabrics and paper, greases and lubricants, paints, polishes, upholstery, textiles, carpets, soil/stain-resistance coatings, mining and oil well surfactants, acid mist suppressants, electronic etching baths, alkaline cleaners, floor polishes, photographic film, and denture cleaners and adhesives. PFAS are also used in food-contact applications for paper protection, including plates, food containers, bags, wraps, and non-food-contact applications (folding cartons, masking papers; Kannan et al., 2004).

The chemical stability and nondegradable nature of PFAS, coupled with their widespread use, has led to global environmental contamination and accumulation of PFAS in aquatic and terrestrial organisms, including humans. PFAS chemicals including, perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA), GenX, perfluorooctyl sulfonamides, and heptafluoro-1-decanol (8:2 FTOH) (Fig. 3) have been detected in environmental matrices and biological matrices (Kannan et al., 2004; Loganathan, 2012, Erickson, 2022a,b). Detectable amounts of PFOS were also found in human blood samples obtained from individuals residing in a number of countries. These findings have raised concerns about environmental contamination by perfluorinated compounds and their possible impacts on ecosystems and human health.

In addition, recent studies showed that chlorinated HOCs such as triclosan and triclocarban that are used as antibacterial compounds in personal care products such as soaps and detergents; as well as other flame retardants such as tetrabromobisphenol A (TBBPA) and hexabromocyclododecane (HBCDs) have emerged as “new HOCs of concern” for the environment and human health (Kemsley, 2014; de Wit et al., 2020) (Fig. 5). Novel flame retardants such as decabromodiphenyl ethane (DBDPE) and hexabromobenzene as well as PFAS replacing compounds such as F53B and ADONA are also “emerging new HOCs of concern” (Fig. 5). Although these chemicals are ubiquitous in the global environment and bioaccumulate in wildlife and humans, their toxic properties are still being investigated even though these compounds seem to affect thyroid hormone homeostasis and cause nervous system and reproductive effects.

2.1. Chlorinated HOCs

2.1.1. Physicochemical properties

Commercially produced HOCs possess unique properties that render them highly persistent in the global environment and cause chronic toxicity to humans and wildlife. PCBs, a well-known chlorinated HOC, are colorless to light yellow with no smell, and are tasteless oily liquids or solids. Some PCBs are volatile and may exist as a vapor in air. The physicochemical properties of PCBs vary widely and depend on the number and positions of chlorine atoms in the biphenyl rings (Kodavanti and Tilson, 1997; Kodavanti et al., 2022b). PCBs have thermal stability and resist both acids and alkalis which made them useful in a wide variety of industrial applications including dielectric fluids in transformers and capacitors, heat transfer fluids, and lubricants. Generally, PCBs are relatively insoluble in water, with solubility decreasing with increasing chlorination. PCBs are readily soluble in nonpolar organic solvents and biological lipids. When burned at high temperatures (i.e., 4500–6500 Btu/lb), the products of PCB combustion include polychlorinated dibenzofuran (PCDFs) and polychlorinated dibenzo-p-dioxins (PCDDs), which are more hazardous than the parent compound. DDT, HCH isomers, chlordane, heptachlor, and toxaphene have properties similar to certain higher-chlorinated and lower-chlorinated PCBs, respectively. Hexachlorobenzene (HCB) is known to build up in grasses, wheat, certain vegetables, and other plants. DDT, on the other hand, tends to bioaccumulate more in higher animals. Heptachlor is a non-systemic stomach and contact insecticide, used primarily against soil insects and termites. It has also been used against cotton insects, grasshoppers, some crop pests and mosquitos to combat malaria. Heptachlor is highly insoluble in water and quite volatile therefore it can be expected to partition into the atmosphere easily. It binds readily to aquatic sediments and bioconcentrate in the fat of living organisms. The half-life of heptachlor in temperate soil is up to 2 years. Toxaphene is yellow waxy solid with chlorine/terpene-like odor and was the most widely used pesticide mixture containing over 670 congeners in the United States until 1975 when it was banned by thirty-seven countries and its use was severely restricted by another eleven. This is a non-systemic, contact insecticide that was used primarily on cotton, cereal grains, fruits, nuts and vegetables. It has also been used to control ticks and mites in livestock. Toxaphene is highly insoluble in water, has a half-life in soil from 100 days up to 12 years, depending on the soil type and climate, and is known to undergo atmospheric transport. Its persistence, combined with a high partition coefficient (log KOW = 3.23–5.50) suggests that toxaphene is likely to bioconcentrate. The physical and chemical stability of the HOCs contributes to their persistence, long range transport, and responsible for human health and environmental problems.

2.1.2. Environmental contamination and exposure to humans

PCBs enter the air, water, and soil during manufacture and use in a variety of applications. PCB containing waste materials in landfills are a potential source of contamination, particularly to the water table secondary to their leakage. These compounds also entered the environment from accidental spills and leaks during the transport of PCB-containing materials like transformers. Once in the environment, PCBs do not readily break down and therefore remain there for a very long period of time. They can easily cycle between air, water, and soil. For example, PCBs can enter the air by evaporation from both water and soil. In air, PCBs can be carried long distances and have been found in snow and seawater in areas such as the Arctic and Antarctic environments, far away from where the PCBs were released into the environment (Loganathan and Lam, 2012; Fig. 4). Because of their persistent and lipophilic properties, PCBs bioaccumulated in various lower trophic organisms (plankton), bivalve mollusks, fish, reptiles, marine mammals, birds, and terrestrial mammals. In concurrence with these findings, these compounds have been detected in fish and other food products (Fig. 4). The primary route of human exposure to PCBs and chlorinated pesticides is/was through consumption of contaminated foods such as dairy products, meat, and freshwater fish. High levels of these compounds were found in human adipose tissues, blood, and milk (Loganathan et al., 1993).

A study in the Netherlands collected breast milk shortly after birth from 209 mothers who intended to breastfeed their newborns. The study also included 209 mothers who did not intend to breastfeed. The neuropsychological function of the children was tested beginning in infancy, with periodic assessment up to 9 years of age. Deficits associated with PCB exposure were observed in both breastfed and non-breastfed infants on several cognitive and behavioral functions (Vreugdenhil et al., 2004). PCB levels in breast milk were based on the four most prevalent congeners, representing about 60% of the total PCBs. The median concentration was 414 ng/g lipid, with a range of 158–969 ng/g. Dividing by 0.6 results in a median of 690 ng/g and a range of 263–1615 ng/g. A similar study in Germany, with equivalent breast milk PCB levels, also documented adverse effects on cognition associated with increased PCB exposure (Winneke et al., 1998).

Chlorinated pesticides such as DDT, HCHs, chlordane, heptachlor, and toxaphene also possess similar properties as some PCBs and as a consequence, these compounds also contaminated the global environment and biota. Chlorinated HOC pesticides also showed a similar pattern of bioaccumulation and biomagnification in the food chain and levels in humans to that observed for PCBs. Human exposure to emerging chlorinated HOCs such as triclosan and triclocarban occurs via directly from soaps and detergents as well as from usage of personal care products.

2.1.3. Effects on human health

The available epidemiological data and studies from laboratory animals provide strong evidence on the toxic health effects of PCBs and chlorinated pesticides. Information on health effects of PCBs is available from studies of people exposed accidentally, by consumption of contaminated rice oil in Japan in 1968 (outbreak of Yusho) and Taiwan in 1979 (the Yu-Cheng incident), by consumption of contaminated fish and meat products and via general environmental exposures. Major symptoms of Yusho disease consisted of acne-form eruptions, pigmentation of the skin, nails, and conjunctiva, increased discharge from the eyes, and numbness of the limbs (Yusho Support Center, 2007). Additional health effects associated with exposure to PCBs in humans and/or animals include liver, thyroid, dermal, and ocular changes, immunological alterations, neurodevelopmental changes, reduced birth weight, reproductive toxicity, and cancer. PCB exposures have been associated with low birth weight and learning and behavioral deficits in children of women who consumed PCB-contaminated fish from Lake Michigan. Furthermore, coplanar PCBs (chlorine substitutions in para and meta positions in the biphenyl molecule) cause dioxin-like toxicity via aryl hydrocarbon receptor (AhR)-mediated toxicity, leading to the development of cancer. In contrast, non-dioxin-like PCBs seem to exert neurotoxicity through their effects on thyroid hormones, neurotransmitters, and intracellular signaling processes (Kodavanti and Tilson, 1997). Serious environmental and health problems, especially in birds, were attributed to DDT and its metabolites (DDE, DDD). In addition, 4,4′-DDE, which is a major metabolite of 4,4′-DDT, has been linked to eggshell thinning and diminished reproductive success in a variety of bird species, including peregrines, hawks, gulls, eagles, terns, cormorants, and many other species (Peakall et al., 1973). Abnormalities in male sexual development in humans also have been associated with the exposure of estrogenic chemicals such as DDT. It is known that 4,4′-DDE does not bind to the estrogen receptor but inhibits androgen binding to the androgen receptor, resulting in inhibition of androgen-induced transcriptional activity and an androgenic effect in mammals. Further, new lines of research indicate that chronic dietary exposure to these chlorinated HOCs may also contribute to obesity and type 2 diabetes (Lee et al., 2014). Zhang et al. (2015) reported that TCDF exposure alters the gut microbiomes in ways that may prove to contribute to obesity and other metabolic diseases. Emerging new organochlorines, such as triclosan and triclocarban, have been used as antimicrobial and antifungal compounds since the 1960s. Triclosan has been reported to hinder cardiac and skeletal muscle contraction in mice and fish. Although triclosan has been considered to be nontoxic to mammals, the adverse effects of continuous, long-term and low concentration exposure remain unknown. Epidemiological studies revealed that levels of triclosan in human tissues, urine, plasma, and breast milk correlate with the usage of this antimicrobial agent. This led to concerns about safety and potential toxicity of triclosan in humans, with special emphasis on early development. The Food and Drug Administration (FDA) recently issued a directive banning the use of triclosan in consumer soaps, justifying the move attributed to data gaps on its effectiveness and safety, indicating the need for more studies addressing these chemical-mediated effects on various tissues including the central nervous system (CNS) (Ruszkiewicz et al., 2017). In addition, triclosan has been reported to disrupt signaling of the endocrine system affecting the function of estrogens, androgens, and thyroid hormones (Louis et al., 2013; Kemsley, 2014). Due to their widespread occurrence and possible toxic effects in wildlife and humans, these compounds are under scrutiny for safety concerns (Kemsley, 2014).

2.2. Brominated HOCs

2.2.1. Physicochemical properties

Similar to PCBs and other chlorinated pesticides, brominated flame retardants such as PBBs, PBDEs and HBCDs are synthetic chemicals that do not occur in nature. FireMaster BP-6 was a mixture of PBB congeners containing two to eight bromines. The major constituents were 2,2′,4,4′,5, 5′-hexabromobiphenyl (56%) and 2,2′,3,4,4′,5,5′-heptabromobiphenyl (27%) (Damstra et al., 1982). PBDEs contain two phenyl rings linked by oxygen (thus the designation as ether; Fig. 2). PBDEs are structurally similar to PCBs (Figs. 1 and 2) and are quite resistant to physical, chemical, and biological degradation. Also, PBDEs cause adverse effects similar to PCBs on nervous, immune, and endocrine systems and influence metabolism of chemicals endogenous to the body as well as the metabolism of foreign chemicals. TBBPA and HBCDs (Fig. 5) are also used as flame retardants. HBCDs are used in building insulation, solder paste, recycled plastics and automobile parts. HBCDs are persistent, bioaccumulative and toxic, and have irreversible health effects (Erickson, 2022a, b). Compared to chlorine atoms, bromine atoms are in general lost more easily from the molecule (more reactive), rendering PBDEs susceptible to various types of degradation and metabolism more readily than PCBs.

2.2.2. Environmental contamination and human exposure

PBDE residues have been detected in indoor air, house dust, and foods, making PBDE exposure to humans possible via multiple routes. PBDEs are found in higher levels in house dusts in the United States than Europe. The contents of vacuum cleaner bags were used to assess household dust exposure to PBDEs in a total of 20 U.S. and German homes. This study found that PBDE-47, PBDE-99, and PBDE-209 were present in the highest concentrations and that the U.S. samples were approximately 50 times higher than samples from Germany. High concentrations of PBDEs are also found in sewage sludge, with levels in the United States running 10–100 times higher than those in Europe. Over half of the sewage sludge produced annually in the United States is applied to land as fertilizer (U.S. Environmental Protection Agency, 1999). Thus, application of sewage sludge may represent a source of exposure to humans and wildlife, through direct contact or uptake by plants. For example, the survey done in 2006 of U.S. foods showed PBDEs in infant soy formula and beef products suggesting that plants as well as animals can be contaminated with PBDEs. Significant levels of PBDEs may be found in outdoor air and even in rural locations. In general, levels of PBDEs were about half those of PCBs in early spring. It is currently well established that there may be significant levels of PBDEs in indoor air and dust. PBDE concentrations in indoor films were 15–20 times higher than in outdoor films.

Levels of PBDEs in human tissues, specifically blood, milk, and fat, have increased exponentially since the 1970s in several countries, including the United States, Canada, and Sweden (Schecter et al., 2003, 2005). The doubling time for PBDE levels in the human body is estimated to be 3–5 years. Breast milk levels of PBDEs are decreasing in Sweden, presumably as a result of a decrease in the use of PBDE-containing products, proving that remediation can improve environmental contamination. The European Union and USA have also decreased PBDE use. Levels of PBDEs among individuals in North America, as measured in blood, breast milk, or adipose tissue, are 10–70 times higher than in Europe or Japan (Shechter et al., 2005, Johnson-Restrepo and Villa, 2016). High levels of PBDEs in North America are attributed to the maximum use of penta BDE mixture (>90%) when compared to the rest of the world. Like other lipophilic compounds, PBDEs readily cross the placenta into the fetus. Thus, the opportunity for PBDEs to interfere with developmental processes, producing developmental effects, is a serious concern. Widespread distribution of TBBPA and HBCDs are also evident, and these compounds have been described as endocrine disruptors and immuno-toxicants, interfering with both estrogens and androgens (Shaw et al., 2012; Johnson-Restrepo and Villa, 2016).

2.2.3. Effects on human health

There have been a number of epidemiological studies on the health effects of environmental exposure to brominated HOCs. PBBs have been reported to cause several adverse health effects in animals as well as humans (Di Carlo et al., 1978). The reports include low birth weight, cryptorchidism, decreased sperm quality, increased incidence of diabetes, altered thyroid hormone levels, and reduced IQ points in children exposed to PBDEs during their development. These reports in humans are supported by studies in animals documenting these adverse effects. There are studies reporting thyroid and androgen effects at doses which do not cause overt toxicity. Extrapolation of a dose in a rodent (rat or mouse) to a comparable dose (or exposure) in humans requires either extensive pharmacokinetic data in both rodents and humans, or the incorporation of a series of assumptions into risk assessment models. Reliance on a series of assumptions obviously introduces substantial uncertainty. One approach that may be useful to circumvent the difficulties associated with rodent-to-human extrapolation is to compare current observed levels of PBDEs in humans to the levels of PCBs that are known to produce adverse effects in humans. One of the most sensitive (perhaps the most sensitive) endpoints for adverse PCB effects is developmental neurotoxicity. It appears that this may be the case for PBDEs as well, based on the available data on structural similarities and adverse effects seen in in vitro as well as in in vivo in animal models. Also, the mechanisms of action for PBDE and PCB neurotoxicity may be the same, as the limited data available suggests that PCBs and PBDEs exert approximately equipotent effects on neuronal intracellular signaling (Kodavanti et al., 2005).

Milk concentrations of PBDEs in the United States are currently well below the levels reported in the Dutch and German PCB studies (see above). The median PBDE concentration in the Texas study was 34 ng/g lipid, with a range of 6–419 ng/g (Schecter et al., 2003). In other studies, the median for breast milk levels was 58 ng/g, with a range of 9.5–1078 ng/g. The median value in breast milk is approximately ten times higher for PCBs than for current concentrations of PBDEs; however, the concentration ranges overlap. The current doubling time for levels of PBDEs in milk in North America is as short as 2.6 years. Assuming that the current exponential increase in human tissues (Schecter et al., 2005) continues, doubling times will continue to decrease. In this case, levels that are known to produce developmental neurotoxicity would be reached in less than 10 years. It is also important to remember that a no-effect level for PCBs has not been determined. Therefore, assuming that PBDEs have effects at approximately the same concentrations as PCBs, it is not known whether current concentrations of PBDEs in the environment are producing adverse effects. In addition, it is likely that the effects of PCBs and PBDEs are additive, at least for neurotoxicity. Therefore, in combination with environmental PCB concentrations, the levels of PBDEs currently in the environment may well be producing adverse effects. Like PCBs, PBDEs also disrupt thyroid hormone homeostasis (Kodavanti et al., 2022a).

Aside from neurotoxic and thyroid hormone effects, other flame retardants such as TBBPA and HBCDs are known to interfere with both estrogens and androgens. TBBPA is also structurally similar to the thyroid hormone, thyroxine and may have endocrine effects. In addition, TBBPA has been reported to suppress immune response by inhibiting expression of CD-25 receptors on T-cells, preventing their activation, and reducing human natural killer cell function.

2.3. Fluorinated HOCs

2.3.1. Physicochemical properties

PFAS is a general term for per- and polyfluoroalkyl substances which are man-made chemicals that have been in large scale production since the 1940s (Fig. 3). Most fluorinated HOCs possess amphiphilic (ionic and neutral) properties. Owing to their thermodynamically strong covalent C–F bonds, these compounds were initially considered nonmetabolizable and nontoxic. To further complicate the situation, many perfluorinated compounds easily isomerize. Because of the physicochemical and biochemical properties such as vapor pressure, water solubility, lipophobicity, metabolic degradability, particle affinity, etc., of HOCs, various perfluorinated isomers will vary significantly with regard to environmental distribution, persistence, and toxicity. Isomer-specific analysis is therefore needed to address the environmental risk posed by these compounds. Perfluorinated compounds are water-soluble in the several parts per million (ppm) range. PFAS with unique surface modification properties readily bind to proteins such as blood globulins.

2.3.2. Environmental contamination and human exposure

Per- and polyfluoroalkyl substances have been produced and used in a wide variety of industrial and consumer products for more than five decades. Because of their widespread use and recalcitrant properties, PFAS have been detected in ice, water, sediment, sewage sludge, etc., and in aquatic organisms, terrestrial and marine animals, and humans. To limit their exposure and detrimental impact on human health and the environment, the United States Environmental Protection Agency (U.S. Environmental Protection Agency, 1999, Federal Register, 2022) has issued health advisory limits for GenX chemicals and PFBS (potassium perfluorobutane sulfonate) are 10 parts per trillion (ppt) and 2000 ppt, respectively. The interim updated health advisories for PFOA and PFOS in drinking water are 0.004 ppt (4 pg/L) and 0.02 ppt (20 pg/L), respectively. Recently, the US Congress is limiting industrial releases of per- and polyfluoroalkyl substances to waterways and public sewage treatment plants. The House of Representatives passed a bill on July 14, 2022, saying that limits of PFAS chemicals in wastewater discharges would apply to producers of organic chemicals, plastics, and synthetic fibers in 2024 (Hogue, 2022). In addition, California’s new laws, AB 1200, ban the use of PFAS in paper, paperboard, and other plant-based food packaging as of January 1, 2023. California joins Connecticut, Maine, Minnesota, New York, Vermont, and Washington in prohibiting PFAS in food packaging (Hogue, 2021). In humans, PFAS compounds were first detected in blood in the 1960s in North America. Later, PFAS have been detected in human blood samples throughout the world (reviewed by Houde et al., 2006). The major routes of PFAS exposure in humans happens via diet such as consumption of contaminated food (mainly seafood) and water as well as inhalation of indoor’s dust/air, and through contact/absorption of other contaminated media. Among other possibilities, consumer products such as nonstick cookware and popcorn bags may be one source of exposure. A study of human blood samples from around the world found the highest levels of PFAS in the United States and Poland, intermediate levels in Korea, Belgium, Malaysia, Brazil, Italy, and Colombia, and the lowest in India (Kannan et al., 2004). Please see chapter 10 in this volume for additional information as it deals with regulatory aspect of these HOCs.

2.3.3. Effects on human health

There have been epidemiological studies on the health effects of environmental exposure to PFAS and these effects are supported by studies in animals indicating adverse health effects associated with PFAS exposures (Lau, 2015). Exposure to some types of PFAS are linked to cancer, developmental problems, hormone disruption, and interference with vaccine effectiveness (Hogue, 2021). During PFOS production, employees in facilities in Decatur, Alabama and Antwerp, Belgium showed no alterations in serum hepatic enzymes, cholesterol, or lipoproteins when serum PFAS levels were less than 6 ppm. However, when PFAS levels exceeded 6 ppm in the year 1995 (0.0–12.83 ppm) and in 1997 (0.1–9.93 ppm) (Olsen et al., 1999), evidence of a peroxisome proliferating effect was seen in individuals with higher serum concentrations. Humans exhibit a weak response to the peroxisome proliferating PFAS, which is in part due to a relatively low level of peroxisome proliferator-activated receptor (PPAR) alpha expression in human liver. This may explain the lack of effect at lower PFOS serum levels. In addition, PFOA was reported to have modulated hepatic responses to obesity and alcohol consumption among the production workers. This was supported by a study in which PFOS/PFOA concentrates in liver and serum, which caused an increase in serum cholesterol in humans following cumulative exposures.

In humans, PFAS can alter expression or secretion of a differentiated gene product (IgM), suggesting an effect on the immune system or perhaps a response to stress/damage. At high exposure levels, both chemicals (PFOS and PFOA) increased solubilization of proteins from lymphoblastic cell lines. Another PFOA-related modulation in hepatic responses is obesity. The cross-sectional study in occupationally PFOA-exposed humans showed 10% increase in mean estradiol levels supporting an endocrine effect. The PFOA induced apoptosis in human HepG2 cells involve production of reactive oxygen species (ROS), mitochondria, and caspase-9 activation (Abudayyak et al., 2021). Patients with a blood concentration over 20 ng/mL of seven widely detected PFAS combined, are at the highest risk of adverse health effects; those with between 2 and 20 ng/mL face some risk. People exposed to high levels of PFAS at work or home should receive regular blood testing and monitoring for adverse health effects, according to a recent report from the National Academies of Sciences, Engineering, and Medicine (Widener, 2022).

3. Summary and conclusions

Increasing levels of chlorinated, brominated, and fluorinated HOCs in the environmental sectors (air, water, soil, sediment) and in human tissues including adipose tissue, breast milk, and placenta continue to be a cause of ecological and human health concern. For HOCs such as PCBs, perfluorinated chemicals (PFAS), and dioxins, consumption of contaminated fish and other food materials has been the main route of exposure to humans. A recent study suggests that in addition to other adverse health effects, chronic dietary exposure to HOCs may also contribute to obesity and type 2 diabetes (Lee et al. 2014, Zhang et al., 2015). The World Health Organization estimated 1.5 billion adults worldwide are overweight or obese and the number of type-2 diabetes increased from 153 to 347 million between 1980 and 2008. Attina et al. (2016) estimated the human cost of long-term low level HOCs could cost the United States $340 billion per annum in terms of health care spending and lost wages. A typical example: PBDE flame retardant exposure leads to IQ point loss and intellectual disability in 43,000 cases with 11 million IQ points loss costing an $266 billion dollars annually. The exposure pathway for these HOCs in humans is mainly from the outdoor environment and consumption of contaminated food. However, for the emerging environmental pollutants such as triclosan, triclocarban, and brominated flame retardants (PBDEs, TBBPA, HBCD), the human exposure pathway is predominantly from indoor contamination. For instance, the widespread use of PBDEs in household items contributes to indoor contamination and is a significant source of human exposure. Similarly, perfluorinated compounds are used in a variety of consumer products and contact with the indoor environment and consumption of PFAS-contaminated foodstuffs are major sources of their exposures. The human exposures and health effects by these HOCs will have a long-term impact, even after a ban on their production, as has been shown for chlorinated HOCs (Loganathan and Kannan, 1994).

A schematic representation of time perspectives of global environmental contamination and human exposure is shown in Fig. 6. Chlorinated HOCs such as PCBs and pesticides rapidly contaminated the environment and biota during the periods of their use for agricultural and public health purposes. The contamination levels declined after the ban or severe restrictions were placed on the production and use of these compounds in most developed countries. However, developing countries continue to use these inexpensive chemicals for agricultural pest control and to control insects that spread malaria, typhoid, dengue fever, etc. Thus, developing countries form the point source for continued global contamination with chlorinated HOCs. Therefore, future chronic toxic effects in humans and wildlife by these HOCs cannot be ruled out. In contrast, brominated and fluorinated compounds are being produced in large quantities and used globally by both developed and developing countries. These compounds are heavily used in indoor appliances and materials. Human exposure pathways for emerging pollutants such as triclosan, triclocarban, PBDEs, TBBPA, HBCD, and PFASs are direct and intimate. In 2016, the US Food and Drug Administration (FDA) issued a rule banning 19 antimicrobial ingredients including triclosan and triclocarban in over-the-counter consumer antiseptic wash products, and the rule took effect starting September 2017 (FDA, 2016). Considerable data have been amassed on the presence of PBDEs and PFASs in indoor environmental sectors (air, water, dust, lint, clothing, food packaging materials, etc.) and human tissues (blood, breast milk, liver, fetus, etc.). Based on the use pattern, recalcitrant property, bioaccumulation, and biomagnification potential of these chemicals, it can be surmised that environmental contamination, human exposure, and health related adverse effects will continue in both developed and developing countries (Fig. 6). Therefore, efforts are essential to monitor the trends of these HOCs due to their widespread use and distribution, and must take preventive measures to minimize exposure to humans from indoor pollution, and dietary exposure to protect human health from possible long-term health effects caused by the HOCs. One of the ways to accomplish this is to strengthen the restrictions on their use and to modernize technological processes to prevent their release.

Fig. 6.

Schematic representation of global environmental contamination trends of HOCs. Chlorinated compounds shown as broken green lines are decreasing with time while fluorinated chemicals shown as dotted blue line are increasing with time and a slow decline in recent years. Brominated compounds (PBDEs, TBBPA, and HBCDs) shown as thick solid red line indicate a substantial increase in contamination levels and a steady state in recent years. The faded lines for all three classes of HOCs indicate projected contamination levels of these HOCs. Adapted from Kodavanti and Loganathan (2017): Organohalogen pollutants and human health. In: Quah, S.R. and Cockerham, W.C. (eds) “The International Encyclopedia of Public Health” Second edition, vol. 5, pp 359–366, Oxford: Academic Press. http://dx.doi.org/10.1016/B978-0-12-803678-5.00318-0.