Flame Retardants: New and Old Environmental Contaminants

Abstract

Flame retardants are a group of compounds used in a variety of consumer goods to inhibit or retard the spread of flames. Several classes of chemical compounds have such capabilities, however, the persistence of these compounds in the environment and their toxicity are crucial points for a risk assessment. Classes such as polybrominated diphenyl ethers (PBDEs) have already been banned in some parts of the world while they are still permitted and extensively used in other parts of the globe. In the need for substitutes for the toxic compounds used, new structures have been synthesized and suggested by the industry as an alternative and substitutives flame retardants. The objective of this review is to address the classes of compounds used as flame retardants in terms of their toxicity to human or non-human organisms and their persistence in the environment.

1. Introduction

Materials with high carbon and hydrogen content are easily combustible, so most materials used nowadays, including plastics, are flammable. Over the last century, the furniture, electronics, upholstery, and textile industries have increasingly employed synthetic materials, which are also used in the transport sector (cars, airplanes, and trains) and at home. Using safety devices against fire, like flame retardants, is important to prevent these materials from burning and harming the society and the environment [1, 2].

But what are flame retardants? Flame retardants (FRs) are chemical compounds employed as safety devices to prevent fires from starting/spreading or to delay ignition, thereby reducing combustible material flammability, increasing escape time, and providing safety to humans and properties [3, 4]. The term “flame retardants” refers to the chemical compound action and not to the compound itself [5]. Various chemical compounds with different physicochemical properties and molecular structures can act as FRs. They can be added to (additive FRs) or incorporated into (reactive FRs) combustible materials, such as wood, plastics, kitchen utensils, appliances, computers, electrical cables, construction materials, textiles, and upholstery [6].

The global FR market is expected to reach about US$53 billion by 2024. In 2019, the world FR consumption amounted to over 2.4 million tons, corresponding to 4.9% growth in market size [1, 7, 8]. China is the largest FR consumer—it accounts for 26% of the global consumption, followed by Western Europe (23%), North America (22%), Asia (18%), and Japan (6%). Together, Central/Eastern Europe, Central/South America, and Middle East/Africa add up to 5% of the world's consumption. Over 175 chemicals are listed as FRs. They are classified on the basis of their chemical composition, but a single compound, aluminum trihydroxide (Al(OH)₃), tops the list as the most consumed FR in the world, corresponding to 38% of all the FRs consumed worldwide. Halogenated flame retardants (HFRs) come next (21%, being 17% brominated FRs and 4% chlorinated FRs), followed by organophosphorus (18%). Other classes like metal-based FRs amount to 14% of the global consumption, followed by FRs based on antimony oxides (9%) [7, 8]. Figure 1 summarizes the consumption of flame retardants.

Figure 1.

Global consumption of flame retardants and their consumption by classes. (Designed using GraphPad prism 8.0.2. Adapted form FlameRetardant-online [7]).

Despite the recent increase in FR use, the first reports on their application date back to 450 BC., when Egyptians employed aluminum to reduce wood flammability. Reports dating back to 200 BC. describe that the Roman civilization used aluminum with vinegar to decrease wood flammability [9]. In modern times, specifically in 1929, polychlorinated biphenyls (PCBs), the first class of FRs, were introduced in the United States of America to meet the need of the electrical industry for an insulator that could act as FR. Later, Europe and Japan also started to produce PCBs. After 37 years, PCB presence in the environment was reported for the first time: a Swedish biologist detected PCBs in fish. Two years later (1968) in Japan, about 1000 Japanese were intoxicated with rising oil contaminated with PCBs. PCBs were widely applied until the 1970s. Then, they were banned in Japan in 1972, and North America stopped producing them in 1976 [1, 10]. However, PCB presence in the environment is still relevant because they are Persistent Organic Pollutants (POPs) with the ability to bioaccumulate and biomagnify, consequently presenting high toxic potential [9, 11].

After PCBs were banned, brominated flame retardants (BFRs) emerged as an economically viable alternative to replace them. Although BFRs and PCBs differ because they belong to distinct chemical classes, the BRF mechanism of action resembles the PCB mechanism of action. In the gas phase, brominated and chlorinated FRs inhibit the combustion process of root chain reaction. HFRs neutralize the high-energy OH˙ and H˙ radicals originating from a chain reaction in fire [12, 13]. However, concerns about HFR toxicity have been raised because they may leach into the environment, with high HFR concentrations being recorded in fish and marine mammals. Concerns about BFR toxic and ecotoxic effects, mainly their carcinogenic and endocrine-disrupting actions in humans, have pressed authorities to legislate about or even ban some BFRs. For example, commercial mixtures of polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecane (HBCD) have been banned or phased out in North America and the European Union (E.U.) [14]. With the new legislation regarding BFRs, the global market has sought economically viable and environmentally friendly alternatives that act similarly to banned FRs. In this context, phosphorus flame retardants (PFRs) have emerged as suitable alternatives for BFRs although they have already been employed for over 150 years [4, 6].

Concerns about FRs being present in the environment grow every day. FRs may easily spread to environmental compartments (air, water, soil, sediments, and even house dust) through dissolution, volatilization, and attrition [4]. Improperly disposed electronic waste and furniture contribute to FR presence in the environment. Weak chemical interaction between manufactured products and FRs applied to them aggravates FR dispersion in the environment, not to mention that numerous compounds employed as FRs have serious effects on human health and the environment. Therefore, ensuring conscious use of these chemicals is crucial [14, 15, 16].

2. Brominated flame retardants

Despite laws discouraging or even banning BRF use, their global production continues to grow because new BFRs are being introduced to replace the banned or phased-out ones. About 75 different BFRs are marketed today, and most of them are found in the environment [12, 14]. The main classes of BFRs are polybrominated biphenyls (PBBs), tetrabromobisphenol A (TBBPA), HBCD, and PBDEs. PBDEs and HBCD are the most employed BFRs worldwide [12, 17]. Table 1 shows the BFR chemical structures and physicochemical properties.

After some poisoning episodes and toxicity evidence involving BFRs, some countries started regulating their use in the same way they regulate the use of PBBs and the commercial PBDE mixture. In the early 1970s, accidental poisoning with PBBs occurred on Michigan Farms—animal feed contamination during production resulted in about 5 million eggs and 15.5 tons of milk products being contaminated and 1.5 million chicken, 30,000 cattle, 6000 hogs, and 1500 sheep dying. After this episode, PBBs were removed from the U.S. market, and this class was banned in the U.S. in 1973 [14, 17]. In the late 1980s, the development of analytical methods helped scientists to begin gathering data about FRs in Europe, North America, and Japan, and environmental and human health concerns started to increase when PBDE presence was reported in human milk [18] and marine animals [19] and rising PBDE levels were identified in environmental compartments including sediments, sewage sludge, the aquatic environment, and biological samples (fish, aquatic birds, and human tissues) [20].

Given the toxicological concern and the fact that POPs represent (some BFRs are included as POP) a growing threat to human health and the environment, in 1995, the council of the United Nations Environment Program (UNEP) requested an international process for evaluating an initial list of 12 POPs, and UNEP asked the Intergovernmental Forum on Chemical Safety (IFSC) to recommend international action on these pollutants. Thereafter, a negotiation process began; the Stockholm Convention on Persistent Organic Pollutants was created and adopted in 2001; and the Convention came into force three years later when 50 countries ratified it. Annex A lists PBDEs, PBBs, and HBCD as POPs to be eliminated [11]. In 2002, the European Union (E.U.) banned PBDEs and HBCD production, followed by the development of framework and directives such as the creation of the Registration, Evaluation, Authorization, and Restriction of Chemicals (REACH), the Restriction of Hazardous Substances, and the Waste Electrical and Electronic Equipment directives [14]. Other initiatives that aimed to reduce and eliminate the two main commercial PBDE mixtures (Penta- and Octa-diphenyl ether – Penta/OctaBDE) were also undertaken in North America (the U.S. and Canada). In 2004, several U.S. states prohibited the use of these two PBDE mixtures in some products. Canada also supported the virtual DecaBDE voluntary elimination by 2013. In 2017, the U.S. Consumer Product Safety Commission (CPSC) petitioned to restrict HFR use as additives and nonpolymeric constituents in electronics, furniture, and children’s products [9, 14]. China added DecaBDE and HBCD to the priority list of substances, which implied restricted production or limited discharges. Taiwan and Japan follow similar examples: they have restrictions on PBDEs and HBCD. On the other hand, although Brazil and India are signatories of the Stockholm Convention on Persistent Organic Pollutants, no comprehensive legislation for FRs exists in these countries [14]. Figure 2 shows a schematic timeline of the regulation of halogenated flame retardants.

Figure 2.

Schematic timeline of regulation of halogenated flame retardant (HFRs). PCBs = polychlorinated biphenyls; PBBs = polybrominated biphenyls; POPs = persistent organic pollutants; HBCD = hexabromocyclododecane; U.S. = United States; E.U. = European Union.

BFRs are divided into three subgroups—additive, reactive, and polymeric—depending on how they are incorporated into manufactured materials. The incorporation mode directly influences BRF presence in the environment. Additive BFRs, such as PBDEs and HBCD, are just mixed at the time of manufacture, so they interact weakly with materials and easily leak into the environment. In contrast, reactive and polymeric BFRs establish a chemical interaction with materials, giving rise to a more stable interaction that results in less BRF bioavailability. Nevertheless, these BFRs should not be neglected because they may be lost to the environment during production or transport [21].

BFRs are ubiquitous in the environment. During the last decades, they have been detected in environmental samples even in places located far away from where they are produced or used (e.g., in the Artic) [22, 23, 24]. Once BFRs are released, they tend to persist, bioaccumulate, and biomagnify, and their physicochemical properties, mainly lipophilicity, may underlie potential toxic effects on the environment [24, 25].

2.1 Environmental occurrence and (eco)toxicological effects

2.1.1 Polybrominated diphenyl ethers (PBDEs)

PBDEs are sold as a mixture of congeners and have three commercial presentations: PentaBDE (pentabromodiphenyl ether), OctaBDE (octabromodiphenyl ether), and DecaBDE (decabromodiphenyl ether). The mixture name refers to the main congeners that compose it. Each PBDE congener varies in the number of bromine atoms and the arrangement of these substituted atoms in the aromatic ring. This gives 209 possible congeners, divided into ten groups: mono-, di-, tri-, tetra-, penta-, hexa-, hepta, octa-, nona-, and decabromo diphenyl ether. The number of isomers in these groups may be 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1, respectively [26, 27]. Depending on the type of material, each mixture has a specific application. For example, DecaBDE is used in diverse polymeric materials, while Penta and OctaBDE are applied mainly in the textile and polyurethane foam industries [26].

PBDEs are released into the environment in different ways. First, they may be released during their industrial production. Second, materials containing PBDEs may release them. Third, goods with PBDEs in their composition may be inappropriately discharged. The latter situation is one of the main sources of environmental contamination with PBDEs. Other sources of exposure to PBDE congeners include use and recycling of products containing PBDEs, such as computers, household appliances in general, upholstery, and furniture [12, 27]. The PBDE physicochemical characteristics, including their high lipophilicity, hydrophobicity, low vapor pressure, and high affinity for particles, contribute to their presence in sediments in ambient compartments, particulate matter in the air, and foods. PBDEs are absorbed by inhalation of domestic and industrial dust, via the dermal route, and even by ingestion of contaminated food, which is aggravated by their ability to biomagnify in the food chain [17, 28].

Regarding the PBDE toxicological aspects, several studies have shown their high toxic potential. Their main effects include hepatotoxicity, neurotoxicity, immunological and endocrine alterations, and carcinogenicity. However, the mechanisms through which PBDEs exert their toxic action are not understood [17, 28]. PBDEs have been detected in human samples, especially blood and breast milk. The latter presentation is particularly alarming. Numerous studies involving human breast milk samples have reported different PBDE concentrations in all the analyzed samples, with the congeners BDE-47, -99, -100, and -153 being the most abundant and frequent [29, 30, 31].

PBDEs, mainly those with lower molecular weight, are structurally similar to thyroid hormones. Therefore, they may disrupt the endocrine system by interfering with hypothalamicpituitary-thyroid axis homeostasis [12, 32]. BDE-71 and -79 decrease thyroid hormone serum levels and induce liver enzyme biotransformation, as shown in studies carried out with mice and rats [33]. Moreover, many PBDE congeners damage mitochondria, increasing reactive oxygen species (ROS) production and oxidative stress, exerting genotoxicity, and inducing apoptotic cell death in isolated rat mitochondrial and hepatocarcinoma cells (HepG2) [34, 35, 36].

2.1.2 Hexabromocyclododecane (HBCD)

HBCD is a high-molecular-weight nonaromatic brominated cyclic alkane with six pairs of enantiomers. It is mainly used as an additive FR in thermoplastics for final application in styrene resins. Being an additive FR, HBCD is easily released into the environment. It has high lipophilicity (log K_ow = 5.6) and low solubility in water (0.0034 mg/l) [12, 37]. Due to these characteristics, HBCD is persistent, with a half-life of 3 days in the air and 2025 days in water. It bioaccumulates with a bioconcentration factor of approximately 18,100 in fathead minnows [38, 39]. Its commercial formulation consists of three isoforms: γ-HBCD (75–89%), α-HBCD (10–13%), and β-HBCD (1–12%) [37]. Enantiomers may behave differently in the environment; for example, γ-HBCD tends to be more toxic than α-HBCD, but α-HBCD is the enantiomer that occurs more often in environmental samples [40, 41].

HBCD has been measured in several environmental compartments, including air and dust, sediments, soil, and sewage sludge, and biological samples (aquatic organisms, marine mammals, birds, plants, and even human samples). In animals, HBCD tends to accumulate in lipid-rich organs, such as the liver, gonads, muscle, and adipose tissue [17, 37, 41].

HBCD presents high toxic potential. HBCD increases catalase transcription because this FR raises ROS concentration. Exposure to HBCD alters a protein involved in the mollusk immune system [42]. HBCD may lead to cellular apoptosis near the heart area, and zebrafish exposure to this compound induces cardiac hypertrophy and arrhythmia [43]. HBCD disrupts the endocrine system in Wistar rats and causes neuro- and hepatotoxicity in mice [44, 45]. HBCD induces cytotoxicity in human hepatocarcinoma cells (HepG2) and human neuroblastoma cells (SH-SY5Y), reducing cell viability. HBCD interferes with T4 metabolism in HepG2 cells and affects the estrogenic activity in breast cancer cells (MCF-7) [46, 47, 48], suggesting that it is hepatotoxic and neurotoxic and that it acts as an endocrine disruptor.

2.1.3 Tetrabromobisphenol A (TBBPA)

TBBPA is a reactive FR and one of the most prevalent FRs in the world. It is mainly applied in the epoxy resin used to produce printed circuit boards. It forms a covalent bond with the polymer structure, so it is less likely to be released into the environment. It is highly lipophilic (log K_ow = 4.75) and has low solubility in water (0.718 mg/l) [17, 37].

Although TBBPA release into the environment is more difficult, it may be released during production, processing, and final usage and disposal of the product it is incorporated into [49]. TBBPA is currently found in many kinds of abiotic and biotic matrixes, and it has been detected in air, water, soil, indoor dust sewage sludge, and sediments [37, 50, 51]. TBBPA may also accumulate in the food chain [52]. Indoor dust is the environmental matrix where TBBPA accumulates the most. This is a concern because human beings spend a considerable time indoors, which increases the risk of adverse effects [37].

Long exposure to TBBPA used to be considered harmless because it was believed to lie below levels that would produce a toxic effect [37, 53, 54]. However, several research studies have demonstrated that even low TBBPA doses disrupt the endocrine system, thyroid hormones, and neurobehavioral functions [37, 55]. Studies on animals have shown some toxic effects. Even at low doses, TBBPA induces toxicity in zebrafish, fathead minnow, and rainbow trout after exposure for 96 h (LC₅₀ = 1.3 mg/l) [56]. In rodents, TBBPA induces nephrotoxicity, oxidative stress in the kidney, increased liver volume, hepatocyte necrosis, and endocrine disruption, but no neurological effects have been found in rats exposed to TBBA [37, 54].

The impacts of TBBPA on human health remain unclear. In vitro studies on human cells have suggested that TBBPA has the toxic potential [37, 55]. TBBPA increases caspase-3 activities and ROS generation; damages mitochondria; induces pathogenesis of several lung diseases in airway epithelial cells (A549) [57]; interferes with immune cell action [58]; induces liver cancer disorder promoting metastasis of liver cancer cells; promotes lysosome exocytosis; and decreases intracellular levels of hexosaminidase (HEXB), cathepsin B (CTSB), cathepsin D (CTSD), and lysosomal enzymes in HepG2 cells [59].

2.1.4 Novel brominated flame retardants (NBFRs)

The ban on some of the most widely used BFRs (PBDEs, HBCD, and TBBPA) has caused novel (or new) brominated flame retardants (NBFRs) to emerge. NBFR production and use have increased in the last decade [60]. Although NBFRs is applied as an alternative to replace traditional BFRs, they share a similar chemical structure with halogenic substitution in a cyclic hydrocarbon/aromatic hydrocarbon, so their physicochemical properties are generally analog. Some NBFRs are hydrophobic, semivolatile, relatively highly lipophilic, and little water-soluble [61, 62]. Table 2 summarizes the NBFR chemical structure and physicochemical properties.

Table 1.

Chemical structure of the main classes BFRs and their physicochemical properties [12, 17].

Source: Pubchem. Chemical structures were designed by the authors by using the software ACD/ChemSketch^®.

Table 2.

Chemical structure and physicochemical properties of novel brominated flame retardants.

Adapted: [61]. Source: Pubchem—https://pubchem.ncbi.nlm.nih.gov/. Chemical structures were designed by the authors by using the software ACD/ChemSketch^®.

Information on NBFR toxicity and ecotoxicity is lacking, causing concern and increasing the number of studies about them. Like restricted and banned traditional BFRs, NBFRs are present in various environmental sources, including air, dust, sewage sludge, and sediments, and they also appear in biotic matrixes (human serum, fish, and birds) [62]. For example, BTBPE is usually measured in sediments associated with others BFRs, at a lower concentration than BDE-209, but higher concentrations than other PBDE congeners. Some studies have detected BTPE in household dust [60, 61, 63, 64].

Due to their physicochemical properties, NBFRs bioaccumulate and biomagnify. DBDPE has Bioaccumulation Factor (BAF) between 6.1 and 7.1 in three fish species (Cirrhina molitorella, Tilapia nilotica, and Hypostomus Plecostomus) from a Chinese river [65]. In southern China, the HBB trophic magnification in an aquatic food chain (invertebrates and fish) is about 2.1 [66], whereas PBT and PBEB biomagnify in waterbirds in the same region [67]. Metabolic rates in the organism influence NBFR bioaccumulation and biomagnification [62].

In vitro studies have reported that NBFRs are hepatotoxic. For instance, 0.1 and 0.2 μM DBDPE up-regulates CYP1A4/5 expression [68], modifying rainbow trout hepatocyte activity [69]. Additionally, it interferes in thyroid hormone deiodinase activity in human liver microsomes; its inhibitory concentration (IC₅₀) is 0.16 μM [70]. As for HBB, it activates the aryl hydrocarbon receptor (AhR), but to a lesser extent than 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Furthermore, 100 μM HBB cannot activate the human androgen receptor (AR) in human hepatocarcinoma cells [71]. On the basis of the AMES test, PBEB and PBT have no mutagenic activity [72].

DBDPE does not show acute toxicity in rats or rabbits: the median lethal dose (LD₅₀) is greater than 5000 and 2000 mg/kg of bw, respectively [69]. However, subchronic and chronic exposure of mice and rats to DBDPE disrupt the endocrine system and alter thyroid hormone homeostasis, respectively [73, 74].

Ecotoxicological studies on aquatic organisms have shown acute DBDPE toxicity to water flea: 48 h EC₅₀ is 19 μg/l. Besides that, zebrafish (Danio rerio) exposure to 12.5 and 25 μg/l DBDPE raises hatched larva mortality and reduces the zebrafish egg hatching rate [69]. Assessment of acute TBB and TBPH genotoxicity in fathead minnow revealed DNA damage in liver cells after oral exposure [75]. Exposure to TBB also increases zebrafish embryo mortality and malformation (LC₅₀ = 7.0 mg/l). Zebrafish exposure to 20 mg/l TBPH (highest tested dose) has no adverse effects. Table 3 summarizes the environmental occurrence and biological effects of the main classes of brominated flame retardants.

	PBDEs	HBCD	TBBPA	NBFRs
Environmental occurrence	Sediment; Particulated matter air Domestic and Industrial dust; Food.	Air and Dust; Sediments, Soil, Sewage sludge Biological Samples (aquatic organisms, marine mammals, birds, plants, and human samples).	Air; Water; Soil; Indoor dust; Sewage sludge; Sediments.	Air; Dust; Sewage sludge; Sediments; Human serum; Fish; Birds.
Biological effects	Hepatotoxicity; Neurotoxicity; Immunological alterations; Endocrine Disruption; Carcinogenicity.	In vivo: Increase oxidative stress (Raising ROS concentration ➔ Increase Catalase transcription); Alters immune system of mollusk; Cardiac toxicity in zebrafish; Endocrine disruption in Wistar rats, Neuro and hepaotoxicity in mice. In vitro: Neuro and hepatotoxicity in human cells lines (SH-SY5Y and HepG2 cells, respectively); Affect estrogenic activity in breast cancer cells (MCF-7).	Endocrine disruption; Changes in neurobehavioral functions – even at low doses; In vivo: Induces toxicity in zebrafish, fathead minnow, and rainbow trout after 96 h of exposure; Nephrotoxicity, hepatotoxicity, and endocrine disruption in rodents. In vitro: Increase apoptotic activity, oxidative stress, and mitochondria damage; Induces pathogenesis of several lung diseases in airway epithelial cells (A549); Interferes in immune cells action; Induces carcinogenic effects in liver cells; Alters intracellular levels of lysosomal enzymes.	In vitro: Hepatotoxicity Up regulates CYP1A4/5 expression; Interferes in thyroid hormone activity; Activates the aryl hydrocarbon receptor (AhR) in HepG2 cell line; Mutagenic activity in AMES test. In vivo: Hepatotoxicity in rainbow trout; Endocrine disrupt in subchronic and chronic exposure in mice, and alter thyroid hormone homeostasis in rats; Acute toxicity in zebrafish; Hepatic genotoxicity in fathead minnow.
References	[17, 28]	[17, 37, 41, 44, 45, 46, 47, 48]	[37, 51, 55, 57, 59]	[62, 69, 71, 72, 73, 74, 75]

Table 3.

Summary of environmental occurrence and biological effect of brominated flame retardants classes.

NBFRs used as an alternative to banned BFRs have frequently been detected in countless environmental compartments, with evidence of persistence under natural conditions—their physicochemical properties resemble the properties of banned BFRs. However, evidence of NBFR toxicity is lacking, and there is no legislation about them. Nevertheless, their in vitro and in vivo ecotoxicity to diverse organisms has been verified [22, 60, 62]. Therefore, finding new compounds that act as FRs and which are safer than older FRs like BFRs to human and environmental health is essential. Phosphorus-based flame retardants (PBFRs), which have already been used for over 150 years, could be suitable alternatives for BFRs [6].

3. Phosphorus-based flame retardants: organic compounds

Although several organophosphate FR categories exist, they are usually divided into three main groups, according to their chemical properties: organic, inorganic, and halogenated [6]. Organic organophosphates include phosphorus derivatives, such as phosphate esters, phosphonates, and phosphinate. Bisphenol-A diphenyl phosphate (BPA-DP), diethylphosphinic acid, diphenylcresylphosphate (DCP), melamine polyphosphate, resorcinol-bis-(diphenylphosphate) (RDP), tricresylphosphate (TCP), and triphenyl phosphate (TPhP) are examples of commonly employed organic organophosphates [6, 76].

Among organic organophosphates, Aluminum Diethylphosphinate (ALPi) is a new generation of halogen-free flame retardants. Despite having high performance in fire control, and gaining space in the manufacture of compounds, little is known about its long-term effects, both in the environmental scope and for human health [77].

Apart from being used as FRs, organic organophosphates have plasticizing and anti-foaming properties. Thus, they are present in various industrial and commercial products, such as eletronic equipment, paints, bedding, textiles, and building materials [78].

Smartphones deserve special attention because they represent a large portion of incorrectly discarded electronic waste, being an important contamination source. Zhang et al. [79] analyzed smartphone samples and found TPhP in all of them. This compound underlies environmental contamination.

Concern about the use of these compounds is related to the fact that they are used as additive FRs. Therefore, they are not chemically bound to the materials and products in which they are added, which facilitates their chronic release into the environment [78]. Another issue is that their concentration in the product decreases over time, thereby reducing its FR properties [6].

To understand the chronic consequences related to the use, adverse effects on the environment, and toxicity of these FRs, their underlying mechanisms must be understood. Moreover, understanding the properties of these FRs and how they behave upon contact with living organisms and the ecosystem is crucial. Organic organophosphate FRs are considered emerging pollutants, and they have been widely studied in recent years.

3.1 Physicochemical characteristics

In general, organophosphates have different physicochemical characteristics. These organic compounds usually have reasonable solubility in water, but this varies according to molecular weight.

Organophosphates with low molecular weight are easily found in the aquatic environment, while compounds with high molecular weight are found in general nature. The octanol/water partition coefficient (log K_ow) of organophosphates ranges from −9.8 to 10.6, which means that they are more lipophilic than hydrophilic. However, they have better solubility in water than BFRs [6, 80].

Organophosphate distribution in the air and the environment varies. According to Henrys law, the vapor pressure also varies [6]. Variations in physicochemical characteristics are directly related to variations in biological effects. Thus, clarifying the mechanisms of these compounds is important [81].

BPA-DP is an organic additive FR with boiling point of 680°C and melting point ranging from 41 to 90°C, low solubility in water (∼ 0.4 mg/l), and log K_ow of 4.5. RDP is found in commercial blends with BPA-DP and is also an additive FR. It has a boiling point of 587°C, no melting point, and is even more insoluble in water (∼1.11 × 10⁻⁴ mg/l), with a log K_ow of 7.41. Both compounds are widely used together due to their advantageous characteristics, such as good thermal stability, high efficiency, and low volatility. They are primarily used as flat-screen protection for TVs and other electronics such as monitors and smartphones [6, 82].

Diethylphosphinic acid, on the other hand, has good solubility in water (4.52 × 10⁴ mg/l), log K_ow of 0.68, melting point of −14°C, and lower boiling point (320°C). DCP also has a low boiling point (235°C), melting point of −38°C, but lower solubility in water (0.24 mg/l) and log K_ow of 4.51. Diethylphosphinic acid is an important compound because it is also a product of the thermal decomposition of aluminum diethyl phosphinate (AlPi), another FR that has been gaining visibility [6, 83].

Melamine polyphosphate is a nitrogen-containing FR. It is chemically linked to the polymer molecule, so it is not considered an additive FR [84]. This compound has the boiling point of 558°C, the melting point above 400°C, good solubility in water (around 1 g/l), and log K_ow of −2.3 [6].

TCP is an additive FR. It is an ester of cresols and phosphoric acid, with the boiling point of 439°C, melting point of 77°C, low solubility in water (∼0.36 mg/l), and log K_ow of 5.11. It is widely used in lubricants, hydraulic fluids, and engine oil. All its isomers (ortho-cresol, para-cresol, and meta-cresol) are active, but o-TCP has gained attention because it is related to cases of neuropathy induced by organophosphates [6, 85].

TPhP is one of the most commonly used additive organophosphates in the industry, and it is also the main contaminant in nature. Its boiling point is 370°C, its melting point is 49°C, it is sparingly soluble in water (∼1.9 mg/l), and its log K_ow is 4.59. Because TPhP has hydrophobic properties, it has a great affinity for sediments and soil, which is why it is frequently found in aquatic environments [6, 86]. Organophosphate distribution in air depends mainly on the values of the octanol-air coefficient (log K_oa). Compounds with log K_oa less than 10 are easily found in the gas phase, while higher values are necessary for compounds to be detected in the particulate form and associated with dust. Values for all the compounds are not yet clear. However, TPhP, for example, has a log K_oa of 10.5 and is easily found in many forms in the home environment [87, 88].

ALPi has excellent heat stability, producing less smoke during burning. The ALPi log K_ow is −0.44 and the water solubility is around 1 mg/l, suggesting low hydrophilic properties. Thus, soil and sediments can be the main target of the accumulation of this compound [89].

Organophosphates have relevant physicochemical characteristics for their applications. In general, they are interesting FRs because they decompose at a lower temperature than polymers used in the production of materials. Thus, heating compounds with phosphorus triggers phosphoric acid formation. This acid envelopes the material, protecting it from pyrolysis and preventing toxic gases from being formed [6, 90].

Table 4 summarizes the chemical structure and physicochemical properties of organic organophosphates.

Table 4.

Chemical structure and some physicochemical properties of novel organic organophosphate flame retardants.

Adapted: [6]. Source: Pubchem. Chemical structures were designed by the authors by using the software ACD/ChemSketch^®.

3.2 Environmental occurrence and ecotoxicological effects

Although organic organophosphates have more interesting characteristics and offer more benefits than BFRs, they are often found in the environment (indoor dust, air, water, soil, and sediments) [91, 92, 93].

Because organophosphate particles are just additives and do not fully bind to the material they are incorporated into, they are easily dispersed. This causes them to be absorbed by suspended dust. Thus, indoor dust is an interesting indicator of indoor exposure to industrial chemicals [64].

Huang et al. [81] analyzed indoor dust from Australian homes, to find that TPhP is one of the most common compounds therein. This is expected if we consider the recurrent use of TPhP. The authors also found BPA-DP in abundance. In fact, this compound is used as a substitute for DecaBDE. The authors concluded that many organophosphates are present in the samples, presenting a high risk to human health [81].

Despite being mostly lipo-soluble, some organic organophosphates have good solubility in water. This has led to their detection in drinking water because treatment stations cannot eliminate these compounds effectively. Thus, the occurrence of organic organophosphates in the aquatic environment poses as much risk to human beings as to the aquatic ecosystem [94].

To identify the presence of organophosphates in water, Kim and Kannan [95] analyzed several samples, such as river water, rainwater, sea water, and tap water samples, collected from various locations in New York State. All the samples presented numerous organophosphates. Chlorinated compounds were identified as the most abundant due to their greater hydrophilicity. Among organic compounds, TPhP was found in over 90% of the river water samples.

On the other hand, compounds with greater lipophilicity are easily found in sediments and aquatic organisms, being the cause of bioaccumulation. In sediments, the composition is directly related to adsorption capacity. For example, clay-rich areas favor greater adsorption of these FRs [96].

Although organic compounds, such as TPhP, are found in sediments, halogenated compounds predominate. Lee et al. [97] and Chen et al. [98] analyzed the composition of water and sediments in lakes. In both studies, the halogenated ester tris-(chloroisopropyl) phosphate (TCPP) were found to predominate. Among non-halogenated compounds, tris-(2-butoxyethyl) phosphate (TBEP) was identified in most samples [97, 98].

Bioaccumulation and biomagnification affect biological characteristics negatively. In fish, for example, growth, sex, food, and maternal transfer are impacted, jeopardizing organism development and the ecosystem as a whole [99].

Organophosphate bioaccumulation has been studied since the 1970s. These compounds have been found in different tissues of rodents, fish, and birds, and they may accumulate in gills, kidneys, liver, and muscle tissue [100].

Wang et al. [101] evaluated the bioaccumulation and biomagnification potential of organophosphates in coral reef fish. These authors found several other types of organophosphates, including TCP. In addition to risking environmental health, this phenomenon may harm human health. Contamination of coral reef ecosystems is extremely relevant and further analysis will be needed in the future.

Indeed, water is an important factor in organophosphate distribution. Particles are dispersed in air through the mechanical friction of the materials. These particles may be included in rainwater and be the main cause of river water and soil contamination [102].

Recycling sites also impact soil contamination by organophosphates directly. In these sites, the soil is in direct contact with different types of materials, mainly electronic waste. Open-air storage of these residues determines local soil contamination and release of these compounds into the air [103].

According to Sánchez-Piñero et al. [104], several organic compounds, mainly BPA-DP and TPhP, are found in soil and dust in public places. Although these authors state that the values are within those authorized by legislation, investigating continuous exposure is necessary.

Thus, soil and water contamination are directly related to human health. Food and drinking water contamination and environmental contamination are directly related, as well.

As a reinforcement, the environmental effects caused by ALPi are still limited and not very specific, although the literature describes that bioaccumulation potential for free halogen compounds is low and that it is easily degraded in the environment [6].

3.3 Toxicokinetics

The prevalence of organic organophosphates in living organisms has become a research target. Toxicological research has suggested that chronic exposure may be directly related to damage to the endocrine system, with a high risk of reproduction being impaired. These compounds may also be related to systemic toxicity, affecting other important organs [101].

Unintentional dust ingestion and skin absorption are the routes with the highest organic organophosphate absorption rate. How these compounds are distributed in biological tissues has not been completely elucidated. However, the highest organophosphates levels have been found in the liver, muscle, and gonads [100].

Compounds are biotransformed through metabolization to diesters and hydroxylation by phase I reactions, followed by conjugation with glucuronide and sulfate in phase II reactions. These reactions happen both in humans and other living beings. Biotransformation products are important for monitoring and evaluating exposure to organophosphates: these products are eliminated fast because they are very hydrophilic [105].

Although organophosphate metabolism is relatively fast, even low concentrations of the metabolites may have physiological and endocrine effects. Considering that human exposure to these compounds is high and that multiple exposure pathways exist, the concern is enormous. For instance, children are generally less able to metabolize and to excrete xenobiotics, so these compounds are more toxic to them [106].

3.4 Toxicological effects

Among the potential adverse effects of organic organophosphates, reproductive and neurological effects are highlighted. Although the toxicity mechanisms of these compounds have not been fully elucidated, the consequences of being exposed to them involve cell apoptosis, ROS production, membrane disturbance, and mitochondrial alterations, among others [100].

Neurotoxic effects are concentration-dependent and inhibit DNA synthesis, decreasing the number of cells and altering neural differentiation. Acetylcholinesterase (AChE) inhibition is another reported mechanism. AChE is a widely used marker in neurotoxicity studies. Tris-(1,3-dichloro-2-propyl) phosphate (TDCPP) and TPhP have an affinity for the nervous system and are commonly associated with neurotrophic factor inhibition [107].

Aquatic organisms are often exposed to organophosphates, which tend to be neurotoxic to them. Sun et al. [108] analyzed the neurotoxicity of the halogenated tris-(2-chloroethyl) phosphate (TCEP) and the non-halogenated alkyl tri-n-butyl phosphate (TNBP) in zebrafish (Dario rerio). More specifically, these authors analyzed locomotor behavior, enzymatic activity, and AChE gene transcription. They found that zebrafish exposure to these compounds in early life stages affects locomotor behavior and gene transcription, suggesting that exposure to organophosphates may be relevant in humans, especially in children.

Exposure at developmental stages may also affect cardiac development. Cardiac development comprises several stages for increasing formation during embryogenesis. When the process occurs properly, chamber formation and maturation, septation, and valve formation take place correctly [109].

Using zebrafish as a model species for studies on developmental toxicity is advantageous and has been gaining ground in several areas of toxicology. Alzualde et al. [110] used zebrafish as a model in developmental toxicity assays to test not only the cardiotoxicity but also the neuro- and hepatotoxicity of organophosphates. These authors found that TPhP affects the heartbeat and reduces locomotor activity and hepatic edema. These data are extremely relevant when it comes to human biomonitoring.

Abe et al. [111] evaluated the toxicity of halogen-free flame retardants in zebrafish to trace a toxicity profile. At all concentrations used, ALPi did not show any sublethal or teratogenic effects, suggesting that ALPi may be a good alternative for brominated flame retardants. However, further studies are still needed to support this information [111].

The crustacean Daphnia magna is also a widely used model for toxicity testing. To test the chronic toxicity of ALPi, Waaijers et al. [112] exposed D. magna by 21 for toxicity tests. The toxicity of ALPi increased with the time of exposure, with low acute toxicity and moderate chronic toxicity [112].

TDCPP and TPhP are also potential endocrine disruptors, altering hormone levels and decreasing semen quality in adult men. Furthermore, the concentration of these compounds in house dust has been correlated with decreased sperm concentration, increased prolactin level, and decreased free thyroxin (T4) level [113].

Given that reproductive system integrity also depends on the organism’s redox state, An et al. [114] tested TPhP and TCPP cytotoxicity in HepG2, A549, and Caco-2 cells. In addition to inhibiting cell viability, these compounds increase ROS production, inducing DNA damage and mitochondrial dysfunction. These changes in redox balance may harm steroidogenesis and even estrogen metabolism, being directly related to reproductive changes.

Epidemiological studies on exposure to organophosphates are also gaining ground, especially when it comes to the early stages of development, when organophosphates may have greater consequences [115].

In a cohort study, Castorina et al. [116] evaluated how exposure to organophosphates, mainly TPhP, affected the cognitive or behavioral development of 310 school-age children. The authors monitored exposure by analyzing metabolites present in pregnant women’s urine. They observed decreased intelligence quotient and working memory, associated with an increased level of the urinary metabolite diphenyl phosphate (DPhP).

Although some compounds have been widely used in research, and even though much information is available, the effects of other types of organophosphates remain to be elucidated. For example, toxicity data on BPA-DP and RDP are limited, so their consequences on human health are unknown [82].

Monitoring organic organophosphate metabolites is necessary to assess and control biological exposure, not to mention that these metabolites may play important biological roles in the toxicity of these compounds. In any case, many studies on organic organophosphates are still needed to understand their toxic effects and to reduce exposure to them [117].

4. Phosphorus-based flame retardants: halogenated compounds

Organophosphorus FRs can be divided into nonhalogenated and halogenated. Halogenated organophosphates have chlorinated forms, which are mainly used as FRs in furniture, building materials, textiles, and electronics [118]. According to the 2012 Chemical Data Reporting from the U.S. Environmental Protection Agency (EPA), about 22,700 tons/year of TDCPP, one of the most used organophosphorus FRs, were manufactured or imported by the U.S.A. in 2010 and 2011 [118, 119]. Tris-(chloropropyl)phosphate (TCPP), tris-(2-chloroethyl) phosphate (TCEP), tris-(1,3-dichloro-2-propyl) phosphate (TDCPP), and tetrakis-(2-chlorethyl) dichloroisopentyldiphosphate (V6) are the main halogenated organophosphates.

4.1 Physicochemical characteristics

4.1.1 Tris-(1,3-dichloro-2-propyl) phosphate (TDCPP)

Tris-(1,3-dichloro-2-propyl) phosphate is commonly abbreviated in the literature as TDCPP, TDCP, or TDCIPP. It a viscous colorless liquid with boiling of 457°C, water solubility of 1.5 mg/l, and log K_ow of 3.8 [6, 120, 121].

4.1.2 Tris-(chloropropyl)phosphate (TCPP)

TCPP is a clear, colorless liquid. Its commercial formulation consists of a mixture of tris-(chloroiso-propyl)phosphate (75%) and bis-(1-chloro-2-propyl)-2-chloropropyl-phosphate (15–30%). It has a boiling point of 342°C, the water solubility of 1.6 g/l, and log K_ow of 2.59 [6, 121].

4.1.3 Tris-(2-chloroethyl) phosphate (TCEP)

TCEP has boiling point of 351°C; however, above 220°C, it rapidly decomposes to carbon monoxide, hydrogen chloride, 2-chloroethane, and dichloroethane. It has a water solubility of 7.0 g/l and log K_ow of 1.44 [6, 121].

4.1.4 Tetrakis-(2-chlorethyl) dichloroisopentyldiphosphate (V6)

V6 has a boiling point of 620°C, the water solubility of 2.1 mg/l, and log K_ow of 1.9. In addition, V6 can be used in conjunction with TDCPP and TCPP, but with specific criteria [6, 121, 122].

Table 5 summarizes the chemical structure and physicochemical properties of Halogenated Organophosphates.

Table 5.

Chemical structure and some physicochemical properties of novel halogenated organophosphate flame retardants.

Adapted: [6, 143]. Source: Pubchem. Chemical structures were designed by the authors by using the software ACD/ChemSketch^®.

4.2 Environmental occurrence and ecotoxicological effects

Because organophosphate FRs are not chemically bound to the original material, they are slowly released into the environment by abrasion and volatilization [123, 124]. Consequently, they are widely distributed in indoor and outdoor environments [124]. In addition, dust and air are important means of human exposure to these compounds via skin and breathing [4]. Furthermore, dust may settle into water bodies, contaminate the water environment and affect aquatic organisms [4].

A study investigated organophosphate FR concentrations in air and dust in 63 homes in Canada, the Czech Republic, and the U.S [125]. The highest concentration of halogenated compounds was found in the U.S.—an average of 1440 ng/g TCEP to 4530 ng/g tris-(2,3-dibromopropyl) phosphate (TDBPP), followed by Canada, and the Czech Republic [11]. Regarding air, TCIPP was detected at the highest average concentration: 73.6 ng/m³ in Canada, followed 26.3 ng/m³ in the U.S. and 16.4 ng/m³ in the Czech Republic [125].

A study carried out in Brazil investigated indoor dust concentrations in different places in Araraquara (Brazil) [22]. The authors observed that TDCIPP (up to 61,200 ng/g) was the second most abundant compound in homes and apartments, [12] and the most abundant compound in cars (from 1050 to 1,600,000 ng/g) [22].

Regarding halogenated compounds in outdoor dust, a study investigated FR concentrations in outdoor dust from urban and rural areas in Nanjing (China) [126]. The authors identified halogenated compounds as the most abundant in both the rural (median: 45.9%) and urban (median: 56.8%) areas [126] and TCPP as the most abundant FR in both studied areas [126].

Assessing sediments is important for monitoring aquatic ecosystems and the aquatic environments quality [127, 128]. Halogenated FRs may accumulate in sediments because they have low solubility in water and a relatively high octanolwater partition coefficient (K_ow) [129].

A study analyzed sediment samples from the Bohai and Yellow Seas (China), to detect organophosphate FRs [130]. The authors identified halogenates as being more abundant than non-halogenated FRs [130]. In addition, TCEP was detected at the highest concentration (from 7 to 671 pg./g of dry weight) [130]. Another study evaluated sediments from rivers in Austria and found TCPP as the most abundant compound in the samples, with a maximum concentration of 1300 μg/kg [131].

Thus, evaluating FRs in sediments is important because, if they are present therein, they will be continuously ingested by aquatic organisms and will eventually accumulate in the food chain, thereby being a potential risk to aquatic organisms [4].

Organophosphate FRs are not completely removed during the wastewater treatment process. Besides that, chlorinated FRs (halogenated group) are more difficult to degrade than non-chlorinated ones [132]. Thus, halogenated FRs have the greatest potential to harm water quality and aquatic health [4, 133].

A study conducted in Germany evaluated the surface waters of the Elbe and Rhine rivers, to detect organophosphate FRs in the samples [134]. In the Elbe River, TCPP and TCEP were measured at concentrations between 40 and 250 ng/l and between 5 and 20 ng/l, respectively [134]. On the other hand, the Rhine river contained smaller concentrations of TCPP (75–160 ng/l) and TCEP (12–25 ng/l) [134].

A study evaluated samples from the Santa Clara River in Los Angeles (California, United States), to find halogenated organophosphate FRs as the main compounds, including TCPP (3.3 μg/l), TDCPP (1.4 μg/l), and TCEP (0.81 μg/l) [135].

Other works have evaluated the presence of halogenated compounds in seawater [136]. A study conducted in China analyzed samples from the Yellow Sea and the East China Sea and detected TCEP (134.44 ng/l), TCPP (84.12 ng/l), TDCPP (109.28 ng/l), and TDBPP (96.70 ng/l) [24]. The authors concluded that the source of these compounds is the municipal and industrial effluent of wastewater treatment plants [136].

Another study conducted in China evaluated types of drinking water samples including tap water, filtered drinking water, bottled water, barreled water, and well water in both urban and rural areas in Eastern China [137]. The authors identified TCPP as being more abundant in barreled water (8.04 ng/l) and well water (2.49 ng/l). The authors found TCEP in low amounts in all types of drinking water, making its carcinogenic risk unlikely [137]. The authors concluded that exposure to organophosphate FRs in drinking water in Eastern China poses no risk to human health [137].

Because halogenated FRs are not fully degraded during wastewater treatment, as already discussed above, they may occur in environmental compartments including air, sediments, and water [125, 130, 132, 137]. Therefore, investigating whether these compounds bioaccumulate in organisms is relevant, as it could damage the ecosystem and human health [138]. The appearance of halogenated FRs in mussels has already been reported in a study conducted in Maizuru Bay (Japan) [139], with TDCPP (18 ng/l), TCP (11 ng/l), and TCEP (11 ng/l) being detected. According to the study authors, these concentrations would not be able to promote adverse effects in the organisms [139]. Another study analyzed fish samples collected from the Pearl River in southern China and domestic birds (chicken and ducks) purchased from farmers living in Qingyuan County [28]. TCEP (82.7–4692 ng/g lipid weight in fish and 33.7–162 ng/g lipid weight in the bird) and TCPP (62.7–883 ng/g lipid weight in fish and 3.89–21.4 ng/g lipid weight in the bird) were detected in all the samples [140].

Studying human exposure to aquatic animals that bioaccumulate halogenated compounds is important. A study has evaluated fish, mussels, and breast milk samples [141]. Aquatic organisms were collected from Swedish lakes, and breast milk samples were collected from women in Swedish cities [141]. The authors found that all the samples contained TCPP at concentrations between 170 and 770 ng/g lipid weight (l.w), and they detected TDCPP only in fish, at a concentration between 49 and 140 ng/g l.w [141]. In addition, they identified TCPP (22–82 ng/g l.w) and TDCPP (1.6–5.3 ng/g l.w) in the breast milk samples [141]. The authors concluded that human exposure to organophosphate FRs via fish and human milk ingestion seems to have minor significance compared to the calculated exposure to these compounds in dust and air [141].

4.3 Toxicokinetics

Humans may be exposed to halogenated organophosphates through inhalation or oral or dermal contact; the general population is exposed to these compounds through food and drinking water [142]. Moreover, occupational exposure to these compounds occurs through vapor inhalation and dermal contact [143]. Consumer exposure includes exposure through vapor inhalation, direct skin contact with halogenated organophosphates on the surface of objects, incidental ingestion of air-suspended particulates or resuspended dust, and ingestion via object-to-mouth behavior by children [144, 145, 146].

TCDCPP, TCEP, and TCPP are rapidly absorbed by the oral route of exposure. Furthermore, TDCPP dermal absorption is significant in rats, and TCPP dermal absorption is significant in humans as revealed by in vitro studies [147]. In addition, TCEP is extensively absorbed during nebulized exposure [148]. After halogenated compounds are absorbed, they are distributed throughout the body without specific accumulation in tissue or organs, but TCEP has been reported to be present in breast milk [149].

These compounds are rapidly metabolized during Phase I and Phase II metabolism. TDCPP is metabolized by a combination of hydrolase, MFO (mixed function oxidase), and GST reactions synthesizing glutathione conjugates, so the main metabolite is BDCPP (bis-(1,3-dichloro-2-propyl)phosphate) [147, 150]. TCEP and TCPP are metabolized by hydroxylation possibly by MFO and CYP 450 enzymes conjugated with glucuronic acid [147]. After that, the metabolic products of halogenated organophosphate FRs are rapidly excreted, primarily in the urine [147].

4.4 Toxicological effects

Exposure to halogenated organophosphate compounds may cause some toxic effects in humans. According to Freudenthal and Henrich, chronic exposure to TDCPP causes benign tumors to appear in Sprague–Dawley rats [151]. A study conducted with patients at a Duke Cancer Institute suggested that increased incidence of thyroid cancer may be associated with exposure to TCEP in the home environment [152]. Additionally, under regulation EC 1272/2008, TDCPP is classified as a category 2 carcinogen with hazard statement H351 “suspected of causing cancer” and TCEP is classified as a “potential human carcinogen” by the E.U [145, 153].

In recent years, endocrine disruption effects of halogenated compounds have been observed. Stapleton and Meeker associated TDCPP concentrations in house dust with hormone levels [64]. They analyzed TDCPP in-house dust collected from 50 men recruited through a U.S. infertility clinic, to observe that increased TDCPP was associated with 3% lower free thyroxine concentration and 17% higher prolactin level [64]. Increased prolactin is a positive effect because it serves a number of important functions involving reproduction, metabolism, and angiogenesis [64, 154]. Studies have also shown that exposure to TDCPP causes thyroid endocrine disruption—exposure of female zebrafish to TDCPP decreases T3 and T4 hormone levels, whereas exposure of Sprague–Dawley rats to TDCPP decreases serum thyroid stimulating hormone [155, 156].

Halogenated FRs and organophosphorus pesticides, such as chlorpyrifos, have similar chemical structures, so these FRs might also exert neurotoxic effects like organophosphate pesticides do [157]. Stapleton performed studies on PC12 cells and observed that TDCPP inhibits DNA synthesis and causes high oxidative stress, without adverse effects on cell viability or growth [157]. Moreover, TDCPP promotes differentiation into dopaminergic and cholinergic neurophenotypes, while TCEP and TCPP only promote the cholinergic phenotype [157].

Finally, there are very limited human health and toxicity data available for organophosphate halogenated FRs, so further studies are needed to understand how they affect humans.

5. Phosphorus-based flame retardants: inorganic compounds

Inorganic FRs act mainly in the solid phase, with thermal decomposition through the release of phosphoric acid. This leads to substrate carbonization, resembling phosphorus-containing FRs [158].

The German report “Substituting Environmentally Relevant Flame Retardants: Assessment Fundamentals” analyzed some FRs on the basis of several evaluation criteria, including potential to accumulate in environmental media (occurrences in humans and the environment), chronic and acute toxicity, emission trend (production, use, and waste disposal), fire by-products (smoke density, gas toxicity, and corrosivity, and fire extinguishing water charges, etc.), and concluded that red phosphorus and ammonium polyphosphate are the least problematic FRs [158].

5.1 Red phosphorus (RP)

RP is a very effective FR in many polymer applications is. It is a stable form of the element phosphorus, which has an amorphous structure [158].

RP acts as an FR via a solid-phase mechanism. It forms a rigid layer that prevents flammable material replenishment in the gas phase, reducing fire and decreasing fire gas toxicity [159]. For polyphosphoric acid (the main component of the aforementioned rigid layer) to be formed, oxygen must be supplied by the polymer or another material used as matrix. Therefore, RP is a more efficient FR in materials with high oxygen content (such as cellulose or plastics containing oxygen) [158]. The RP concentration varies from 2 to 10% of the total weight. RP reduces the toxic smoke formation and heat release, preventing large fires from occurring [160]. Its use restrictions are based on color—because it is red, it cannot be used in white or light products given that it is difficult to mask even with dyes, [158].

RP is mainly used in condensation polymers (polyamide, polyester, polyurethane, polyisocyanurate, and epoxy resins), but it can also be used in dispersions for textile finishing, polyamide components for electrical and electronic devices, fiberglass reinforced plastics for electrical applications, synthetic glues, and automotive textiles [158].

According to some studies, the use of RP as FR is not problematic because it does not dissolve in water easily, making the risk of environmental contamination with phosphorus unlikely. Organ intoxication effects are also unlikely, and RP may only cause skin irritation. Therefore, the use of this FR has a low ecological and human health impact, as long as it is not mixed with white or yellow phosphorus [158].

The risks of contaminating the environment with phosphorus as a result of using RP as FR are unlikely. There are no data on RP concentrations in air, soil, or water. Like microencapsulated phosphorus, inert RP does not pose a threat to the environment [158].

The occurrence of phosphorus compounds in environmental samples cannot be analyzed separately from the natural occurrences of phosphorus compounds and cannot be seen as a consequence of the use of RP as FR [158]. Accumulation is hardly a consequence of RP used in plastics because this FR degrades fast, with phosphine and phosphoric acid formation. The effects on aquatic systems are not alarming because phosphorus concentrations are low compared to natural occurrences [158].

RP oral ingestion is unlikely because it is degraded in the environment when it is eliminated in sewage plants through adsorption to sewage sludge. Whether resorption occurs when RP is microencapsulated and ingested orally has not been examined, but resorption is more likely negligible and organ effects are improbable [158].

If inhaled, RP LD50 in rats is 4.3 mg/l, which indicates moderate acute toxicity because metabolism is rapid. Eyes and mucous membranes may become irritated probably due to acid formation. Therefore, affected individuals should not be taken for toxicological evaluation because phosphoric acid may be formed due to the high reactivity of phosphine. Nevertheless, oxides released during a fire should not be ignored because they may irritate the affected persons skin [158].

The RP physicochemical properties prevent it from evaporating at room temperature and from solubilizing in water. Because oxygen is present in the environment, RP forms some phosphates that generate phosphine (PH) through a complex chain of chemical reactions. As toxic as phosphine is (with prescribed exposure limits for humans being TLV/TWA—0.3 ppm or 0.4 mg/m³; TLV/STEL—0.75–1 ppm), it is also very reactive and produces non-toxic phosphates [159].

Some techniques for RP production have been developed and improved, so phosphine is no longer a problem. Phosphine arises if RP is treated at high temperatures or if it is exposed to humidity. If during compounding the RP levels are kept below the TLV/TWA limit, phosphine formation may be drastically reduced, and they can be removed in a well-ventilated area [159]. RP should not be stored in closed containers because phosphine may be formed [158].

RP in the powder form is flammable and has therefore been regulated as a potentially hazardous material. The safest and most ecological way to transport RP is to microencapsulate it. But the newly developed RP grade is highly stable, safer to use, and easier to handle in terms of cleaning and service life [159].

During a fire, RP is easily oxidized to phosphorus oxides that do not form phosphine: in fact, phosphorus oxides remain as polymeric phosphoric acid or phosphates that can be removed in the incinerators’ flue gas treatment system [159].

5.2 Ammonium polyphosphates (APP)

APP is a crystalline inorganic salt with the chemical formula NH₄PO₃. This FR retardant is found in the crystalline or liquor form and contains nitrogen and phosphorus (Figure 3). Like most FRs containing phosphorus, APP begins to decompose at temperatures above 225°C and acts by releasing phosphoric acid and carbonizing [158].

Figure 3.

APP chemical structure.

The composition of ingredients varies depending on the manufacturer. For example, APP 422 from Clariant GmbH, contains 31.5% phosphorus and 14.5% nitrogen [158]. APP is often used in conjunction with other compounds such as polyalcohol as a carbon dioxide dispenser, melamine as a blowing agent, and aluminum trihydrate (ATH) in resins [158].

In intumescent systems, APP is combined with carbonic compound distributors and blowing agents. This mechanism releases non-flammable gases, preventing oxygen from being supplied to the flammable substrate. APP is used in polyurethane foams (hard and flexible), polypropylene, epoxy and polyester resins, cellulose-containing systems, and wash-resistant textile backings [158].

Some studies have reported that, from a toxicological viewpoint, APP is not problematic, or that it has a very low ecological and human health impact. Air contamination caused by APP is unlikely due to its physicochemical properties [158].

There are no data available on APP reabsorption by oral ingestion, but a high rate can be predicted from experiments with similar compounds. APP is metabolized to ammonia and phosphate, which integrate with the nitrogen and phosphate cycle. However, there are no considerable concentrations in the body or toxic effects that need to be feared. If APP comes into contact with skin, it causes some irritating or sensitizing effect because hydrolysis occurs in the presence of acid and ammonia salts in an aqueous medium. The LD50 data (>2000 mg/kg) allows no conclusions about chronic toxicity [158].

Rapid APP decomposition into ammonia and phosphate occurs in soil and sewage sludge, so water eutrophication must be taken into account. However, there are no exact data on the relevance of the volume. During a fire from plastics containing APP, nitrogen oxide and ammonia are formed, as well as phosphorus oxide. These gases are aggressive, and the health effects cannot be ignored [158].

Table 6 summarizes the biological effects caused by organic, halogenated, and inorganic phosphorus-based flame retardants.

Abbreviation	Type	Toxic effect	Reference
TPhP	Organic	Neurotoxic—AChE inhibition; Cardiotoxicity, affecting heartbeat and reduces locomotor activity and hepatic edema; Decreased sperm concentration, increased prolactin level, and decreased free thyroxin (T4) level; Cytotoxicity in HepG2, A549, and Caco-2 cells, increasing ROS production, inducing DNA damage and mitochondrial dysfunction; Affect the cognitive or behavioral development.	[107, 110, 113, 114, 116]
TNBP	Organic	Affect locomotor behavior and gene transcription.	[73]
ALPi	Organic	Any sublethal or teratogenic effects, suggesting that ALPi may be a good alternative for brominated flame retardants; Increase with the time of exposure, with low acute toxicity and moderate chronic toxicity (Daphnia magna).	[111, 113]
BPA-DP	Organic	Toxicity data are limited.	[6]
RDP	Organic	Toxicity data are limited.	[6]
TDCPP	Halogenated	Affinity for the nervous system and are commonly associated with neurotrophic factor inhibition; Altering hormone levels and decreasing semen quality in adult men;	[107, 113]
TCPP	Halogenated	Cytotoxicity in HepG2, A549, and Caco-2 cells. Increasing ROS production, inducing DNA damage and mitochondrial dysfunction.	[114]
TCEP	Halogenated	Inducing thyroid cancer	[152]
RP	Inorganic	Low human health impact as long as it is not mixed with white and yellow phosphorus	[158]
APP	Inorganic	Low human health impact	[158]

Table 6.

Summary of biological effects caused by organic, halogenated, and inorganic inorganic phosphorus-based flame retardants.

6. Conclusion

Fire has always been an enemy of humanity and can cause a lot of destruction. Over time, more and more people live with materials with high flammability, creating the need to control the flammability of these materials. Thus, flame retardants are widely used to reduce the risk of fire, as a safety device. However, other risks can arise with the wide use of this class of compounds, including the risks to environmental and human health.

As flame retardants are being used, research and new knowledge are being generated to understand the behavior of these substances in the environment and the consequences for the ecosystem and human health. Classes of flame retardants begin to be legislated and controlled due to their highly toxic potentials, such as PCBs, PBDEs, and HBCD. On the other hand, new compounds have been introduced to replace the legislated flame retardants, here, phosphorus-based flame retardants emerge as an effective and possibly ecofriendly alternative to the old flame retardants.

As the need arises for alternatives to the old flame retardants, new substances are being introduced into the market with the purpose of reducing the damage that could be caused. However, this insertion occurs without full knowledge of the toxic effects of these new products. Furthermore, over the last decades, we can observe a pattern of substitution of harmful or legislated flame retardants to another that initially was human and ecofriendly but, after few years it turns out and it was harmful. In this way, the environmental consequences can be exacerbated, as can the exposure of ecosystems. In this chapter, we provide a review of the main classes of flame retardants and its possible substitutes, trying to understand the behavior of these substances in the environment and their toxicological consequences for the ecosystem and human health.