Bioaccumulation and Biotransformation of Chlorinated Paraffins: History
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Contributor:
Liujun Chen
,
Bixian Mai
,
Xiaojun Luo

Chlorinated paraffins (CPs), a class of persistent, toxic, and bioaccumulated compounds, have received increasing attention for their environmental occurrence and ecological and human health risks worldwide. Understanding the environmental behavior and fate of CPs faces a huge challenge owing to the extremely complex CP congeners. The bioaccumulation and biotransformation of CPs, including the occurrence of CPs in biota, tissue distribution, biomagnification, and trophic transfer, and biotransformation of CPs in plants, invertebrates, and vertebrates in detail are summarized and integrated. Biota samples collected in China showed higher CP concentrations than other regions, which is consistent with their huge production and usage. The lipid content is the major factor that determines the physical burden of CPs in tissues or organs. Regarding the bioaccumulation of CPs and their influence factors, inconsistent results were obtained.

  • chlorinated paraffin
  • occurrence
  • bioaccumulation
  • biotransformation
  • distribution
  • metabolism

1. Introduction

Chlorinated paraffins (CPs), known as poly-chlorinated n-alkanes (PCAs), are produced by the chlorination of different n-alkanes with the formula CnH2n+2−mClm. They are divided into short-chain (SCCPs, C10–13), medium-chain (MCCPs, C14–17), and long-chain CPs (LCCPs, C ≥ 18) according to the lengths of their carbon chain. Recently, very-short-chain CPs (vSCCPs, chain lengths below 10) have been reported in wildlife and human samples [1][2]. CPs have the advantages of low volatility, flame retardancy, good electrical insulation, and low cost. Consequently, CPs have been used as plasticizers, flame retardants, and additives in metal cutting fluids, adhesives, coatings, rubber, and sealants with a global CP production of more than two million tons per year [3][4].
CPs are highly persistent in the environment, are bioaccumulative and toxic to animals and humans, and have a high probability of undergoing long-range environmental transport. SCCPs were designated as persistent organic pollutants (POPs) and listed in annex A of the UN Stockholm Convention in 2017. This regulation has resulted in a limit in the use and production of SCCPs. A substantial amount of research has reported the occurrence of CPs in abiotic media such as sediment [5], soils [6], and airs [7], as well as in the biota [8]. Nevertheless, the understanding of the environmental behavior, fate, and potential risks is still scarce because of the extremely complex CP congeners and the difficulties regarding accurate quantification.
Some reviews on CPs have been produced in the last ten years, and most of them focused on analysis method, environmental distribution, and toxicity [9][10][11][12]. The data on CPs in organisms have increased over the last decade, and studies on the trophic transfer of CPs in both aquatic and terrestrial food webs has increased substantially in recent years [13][14]. The concerns regarding the biotransformation of CPs have also been continuously increasing.

2. Occurrence of CPs in Organisms

Data on CPs in biota were mostly obtained from China from 2010 onwards. This is plausible considering that China has been the largest global producer and consumer of CPs since 2007, producing up to 1,000,000 metric tons in 2009 [15], and CP contamination has become a major environmental concern in China. A precise comparison of CP concentrations among different regions or countries is difficult since the samples were collected in different periods and the species differed between different studies. Additionally, the concentrations were expressed using different units (ng/g dw, ng/d ww, ng/d lw, etc.). Generally, CP concentrations in aquatic organisms from China were several orders of magnitude higher than those from other regions or countries. Relatively high CP concentrations were found in aquatic organisms collected from e-waste recycling sites such as those in Qingyuan and Taizhou, China [13][16][17]. SCCPs have also been reported at relatively high concentrations in mollusks from the Liaohe estuary [18] and fish from Liaodong Bay [19].
Data on CPs in terrestrial biota are all from China. Similar to aquatic organisms, high SCCP concentrations were also detected in terrestrial biota collected from e-waste recycling areas [16]. Surprisingly, CP concentrations in plants and animals collected from Tibet, China, were also comparable with those in the e-waste recycling area [14][20][21], indicating high CP pollution in this area, although the intensity of anthropogenic activities in this area was thought to be lower.
SCCPs and MCCPs have been detected in human tissues including fingernails, hair, blood, cord serum, the placenta, and breast milk from different countries [22][23][24][25][26][27][28][29][30][31][32][33][34]. Human foodstuffs, including aquatic food, meat, baby food, noodles, green tea, cereals and beans, milk, rice seeds, and oil food, from China and Europe were also found to contain SCCPs and MCCPs [35][36][37][38][39][40][41][42][43][44][45][46][47][48]. Concentrations of SCCPs in the foodstuffs were generally higher than concentrations of MCCPs, with the exception of oil-based vitamin E dietary supplements collected in Germany [48]. However, the concentrations of MCCPs were similar to or more than the concentrations of SCCPs in most of the human tissues samples, indicating a higher bioaccumulation of MCCPs than SCCPs in humans.
Earlier studies focused more on SCCPs in aquatic species; however, research on vSCCP, MCCPs, and LCCPs in terrestrial species or humans has increased in recent years [2][49][50]. LCCPs were first detected in human blood samples in China with median concentrations of 150 ng/g lw, which were lower than those of SCCP (3500 ng/g dw) and MCCPs (740 ng/g dw) [28]. Since then, LCCPs were detected in marine organisms from the Baltic Sea (nd: 130 ng/g dw) [51], Greenland and Iceland (<0.41–930 ng/g dw) [1], aquatic and terrestrial organisms from Germany (nd: 2400 ng/g dw) [50], plants from Beijing, China (27–561 ng/d dw) [52], and cereal and legume samples from China (48.1–664 ng/g dw) [26]. Compared with SCCPs and MCCPs, the concentrations of LCCPs in the abovementioned samples were on the same order of magnitude as those of SCCPs and MCCPs. vSCCPS were initially quantified in aquatic biota samples from the Yangtze River Delta in China with high levels (2.6–8400 ng/g lw) [2] compared to those of vSCCPs in bivalves and marine mammals from Greenland and Iceland (<0.12–34 ng/g lw) [53] and those in biota from Germany (nd: 65 ng/g lw) [50]. vSCCPs were also detected in fish and mosses from Liaodong Bay and the Antarctic at relatively low levels (3.4–153 ng/g dw) [49].

3. Tissue Distribution of CPs in Organisms

Several in vivo exposure experiments investigated the tissue distribution of CPs in organisms [54][55][56]. Lipid content was the main factor in determining the physical burden of CPs in organ tissue. The high concentrations of SCCPs (based on wet weight) were in the abdominal fat and the feces, and the low concentrations were in the blood, meat, and bile fluid when broiler chickens were exposed to SCCP (60% Cl) via diet. An exposure experiment using laying hens obtained similar results [55]. The amount of SCCP deposited in the leg meat was higher than that in breast meat due to the higher fat content of leg meat compared to breast meat in both broiler and laying chickens. An experiment using rat exposure to SCCPs, MCCPs, and LCCPs revealed that approximately 57.0–76.5% of CPs were deposited in the liver, 23.1–42.4% of CPs in the fat, while approximately 0.3–1.4%, 0.01–0.04%, and 0.002–0.022% CPs were in the blood, kidneys, and lungs, respectively [56]. The same dependence on the lipid content was also observed in field fish samples. Sun et al. (2017) [13]. found concentrations of SCCPs (based on wet weight) in snakehead and mud carp in the order liver > gill > kidney > skin > muscle, and the SCCP concentrations in the tissue were positively correlated with the content of lipids in tissues (p < 0.001). However, when the concentration was expressed based on lipid, the muscle SCCP concentrations were the highest among all tissues in the laying chicken [57]. Sun et al. (2020) found similar results in chicken collected from an e-waste recycling site [58]. The concentrations of SCCPs were in the order muscle (6200 ng/g lw) > fat (2400 ng/g lw) > liver (1100 ng/g lw). This tissue distribution was due to the low lipid content in muscle and metabolism of SCCPs in the liver.
The deposition of CPs in tissue was also related to carbon chain and degree of chlorination. Mézière et al. (2021) [57] conducted an experiment using laying hens exposed to different degrees of chlorination (low %Cl and high %Cl) SCCPs, MCCPs, and LCCPs to investigate the accumulation and distribution of CPs in biota after ingestion. All C10–C36 CPs were detected in the liver. However, differences were observed in CP distribution: LCCPs with high %Cl were retained in the liver, while LCCPs with low %Cl circulated through the serum and were distributed in the different compartments but were mostly excreted through the eggs; SCCPs and MCCPs were found in all tissues at similar levels. SCCPs with low %Cl were detected at lower levels compared to SCCPs with high %Cl and MCCP, implying a higher biodegradation potential for SCCPs with low %Cl compared to CPs with higher %Cl. Du et al. (2020) [59] analyzed the vSCCPs, SCCPs, MCCPs, and LCCPs in different tissues of terrestrial short-tailed mamushi (Gloydius brevicaudus) and the semiaquatic red-backed rat snake (Elaphe rufodorsata) from the Yangtze River Delta of China. The tissue distribution of total CPs content (ww) in the two snakes was in the order fat (44–1300 ng/g ww) > muscle (39–550 ng/g ww) > liver (41–490 ng/g ww). vSCCPs (C6–9) and SCCPs (C10–13) were preferentially distributed in the snake liver, while fat was an important storage compartment for MCCPs (C14–17) and LCCPs (C > 18).
The tissue distribution of CPs also exhibited the specific homologues for specific tissues. In the studies of CPs in chicken, Sun et al. (2020) [58] observed that the concentration ratio of muscle to liver of CPs decreased with increased log KOW, while the concentration ratio of fat to liver increased with increasing log KOW. This result indicated that the fat prefers to accumulate high log KOW homologues compared to the liver and muscle. Muscles that preferred to deposit SCCP with lower log KOW compared to liver were also found in fish samples [13]. The difference in the lipid component was responsible for this tissue distribution. The proportions of neutral lipids (triglycerides) to total lipids in the abdominal adipose tissue, liver, and muscle of chickens were 98%, 52%, and 32%, respectively, while the proportions of polar lipids (phospholipids) to lipids were 1.7%, 46%, and 65%, respectively [30][60]. It is likely that SCCPs with higher log KOW have a stronger affinity with neutral lipids than polar lipids.

4. Bioaccumulation of CPs in Organisms

Bioaccumulation is the process of accumulation of chemicals in an organism from the environment through diet, skin absorption, or respiration that occurs when the chemical concentration in the organism exceeds that of the surrounding environment or diet after reaching equilibrium [61]. The bioaccumulation potential of chemicals for evaluation comprises the bioaccumulation factor (BAF), bioconcentration factor (BCF), biomagnification factor (BMF), and trophic magnification factor (TMF). Generally, the chemical is considered to be bioaccumulative when BAF/BCF > 5000, BMF/TMF > 1 [62].
Numerous laboratory and field studies showed that CPs had bioaccumulation potential. The values for the logBAF of ∑SCCPs and ∑MCCPs were estimated as 2.70–4.45 and 2.25–3.64, respectively, for aquatic insects from Qingyuan at an e-waste recycling site [63], 1.48 for soil–vegetation of ∑SCCPs from the Arctic [64], 4.5–6.05 for fish of ∑SCCPs from Liaodong Bay [19][65], and 2.04–3.79 for marine organisms of ∑SCCPs from the East China Sea [66]. Following the administration of 14C-labeled CPs in bivalve blue mussels (Mytilus edulis), the values for the BAF of MCCPs (C16, 34% Cl) and SCCPs (C12, 69% Cl) were estimated at 7000 and 13,900, respectively, indicating that SCCPs with short carbon chains and high chlorination had strong bioaccumulation [67]. The log BCF values of five different CPs (Cereclor S45:MCCP 45% Cl; Cereclor 50LV: SCCP 50% Cl; Huels 70C:SCCP 70% Cl; CP-42: 42% Cl, C10–C17, C21–C31; CP-52: 52% Cl; C9–C29) in Daphnia magna were assessed as 6.7–7.0 (Lkg lipid−1). All the CPs tested were bioaccumulative in D. magna [68].
The BMF of ∑SCCPs for oyster–mangrove crab (2.4) from the Pearl River Estuary [69], Ngas–Agas (1.9) from Antarctica [64], and prey species–finless porpoises (3.1–6.7) and prey species–Indo-Pacific humpback dolphins (11–38) from Hong Kong [70] were all >1, indicating the biomagnification of SCCPs in these marine food chains. In juvenile rainbow trout that were exposed to CPs (C10H15.3Cl6.7, C14H23.3Cl6.7 and C18H31.4Cl6.6), the BMFs for ∑SCCPs ranged from 0.9 to 2.8, demonstrating that these CPs had the potential to biomagnify [71]. In an e-waste recycling pond in South China, the BMF values for ∑SCCPs and ∑MCCPs for fish–water snakes were 2.9 and 2.95, respectively, indicating biomagnification [17]. The mean BMFs of vSCCPs, SCCPs, MCCPs, and LCCPs in the food chain (black-spotted frog–red-backed rat snake) were 2.2, 1.9, 1.8, and 1.7, respectively, indicating the potential for biological magnification. This was the first study to address the biomagnification potential of vSCCPs and LCCPs [59]. On the other hand, biodilution was also reported in both marine and freshwater food chains. For examples, the BMFs of SCCPs and MCCPs for fish–bird in an e-waste recycling pond were 0.08 and 0.1, respectively [17]. The BMFs of ∑SCCP were 0.46 for gammarid–cod from Ny-Ålesund in the Arctic [72], and were 0.282 and 0.212, respectively, for plant–plateau pika, as well pika–eagle from the Tibetan plateau [14].
The varied trophodynamic behavior of CPs was also reported in previous studies. Trophic magnifications of SCCPs were observed in an aquatic food web from Dianshan Lake in Shanghai (TMF: 1.19–1.57) [73], marine food web from the East China Sea (TMF: 3.98) [66], and zooplankton–shrimp–fish food web (TMF: 2.38) [19] and fish food web (2.57) [65] from the Liaodong Bay Sea. The TMF of SCCPs and MCCPs in a marine food web from Hong Kong (TMF: 4.29 and 4.79, respectively) [70] and in a terrestrial food web dominated by insects from an e-waste recycling site in southern China (TMF: 2.08 and 2.45, respectively) indicated trophic magnification [13]. The TMF of SCCPs and MCCPs in the invertebrates–forage fish–lake trout food webs from Lake Michigan and Lake Ontario were 0.97 ± 0.33 and 1.2 ± 0.51, respectively, which is in the margin of trophic magnification and trophic dilution [74]. The lake trout had lower SCCP concentrations than several of their prey on a lipid basis, which did not support the biomagnification of CPs. Meanwhile, the trophic dilution of CPs in food webs was also reported. The TMF of SCCPs in an aquatic food web from an e-waste recycling site was 0.17 [13]; the TMF values of SCCPs and MCCPs in a freshwater food web from France were <1 [75]; the TMF of SCCPs in mollusks from Bohai was 0.396; and the TMF of SCCPs in plant–pike–eagle food chain from the Tibetan plateau was 0.392 [14].
The biomagnification and biodilution of CPs in both aquatic and terrestrial food webs were reported in different regions and in different food webs. The structure of food webs; size of organisms collected; food habits; biotransformation of SCCPs in organisms; treatment method for any outliers; and even environmental parameters such as water temperature, dissolved organic matter, and suspended particles may have all contributed to the differences obtained [19][76][77].
Both the chlorine content and the carbon chain length determine the physicochemical properties of CPs [78]. Therefore, both may affect the bioaccumulation capacity of CPs. In a laboratory exposure experiment, pumpkin seedlings were exposed to four kinds of SCCPs: 1,2,5,6,9,10-C10H16Cl6 (HexCD), 1,1,1,3,8,10,10,10-C10H14Cl8 (OctaCD), 1,1,1,3,10,11-C11H18Cl6 (HexCU), and 1,1,1,3,12,13- C13H22Cl6 (HexCT). The cumulative amounts of OctaCD, HexCT, HexCU, and HexCD in pumpkin were 59.9%, 40.5%, 33.4%, and 23.6% of the parent SCCP, respectively, showing that bioaccumulation increased with the carbon chain length and chlorination degree in plants [79]. The soil–vegetation BAFs of SCCP in the Arctic support this finding. BAFs increased with increasing content carbon and chlorine, and the number of carbon atoms was the primary factor regarding the bioaccumulation of SCCPs [72]. Fisk et al. [71] observed that the BMFs of CP congeners (C10Cl6.7, C14Cl6.6, and C18Cl6.7) in juvenile rainbow trout increase with increasing carbon chain length in a given chlorination content. The BMF of SCCP congeners in a food chain (tapertail anchovy–Bombay duck) collected in the Pearl River Estuary showed significant positive correlations with both number of carbon and chlorine atoms of SCCP congeners [80]. All these results indicate that both the length of carbon chain and the chlorine content affect biomagnification of CPs in organisms.
However, some field monitoring provided inconsistent results. The BMFs of SCCP congeners in the food chain Agas–Ngas from Antarctica decreased from 2.3 for C10 to 0.8 for C13 [64]. Houde et al. (2008) [74] also observed that BMFs in a food web from Lake Michigan decreased with the increasing carbon chain length of SCCP congeners. In aquatic organisms collected from a large sewage treatment plant in Beijing, the BMF of SCCP homologues exhibited significant correlation with chlorine content while no significant correlation was found between the BMF and carbon chain length [76]. The correlation between bioaccumulative potential and chlorine content but not between carbon chain length were also found in other food chains/webs, such as bivalves–sediments from Bohai Sea [81], and fish–water snake, fish–waterbird food [17], and insect-dominated terrestrial food webs in an e-waste site in South China [82]. These results imply that chlorination degree rather than carbon length affects the bioaccumulative potential of CP congeners.
The low bioaccumulative potential with low chlorine content of these CPs may be due to the metabolization of low-chlorine-content CPs, which were more easily compared to CPs with higher chlorine content. The carbon chain length has positive and negative effects on the bioaccumulation of CPs. On the one hand, the lipophilicity of CP congeners increases with the increasing carbon chain length, which is beneficial to bioaccumulation. On the other hand, the molecular sizes of CP congeners increase with the increasing carbon chain length, which reduces the bioaccumulative potential of CP congeners. This could explain the inconsistent results of the observed effects of carbon chain on the bioaccumulation of CP congeners.

This entry is adapted from the peer-reviewed paper 10.3390/toxics10120778

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