a review of phthalates and the associated reproductive and decelopmental toxicity towards fish msc...
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A review of phthalates and the associated reproductive and developmental toxicity towards fish.
Masters literature thesis -‐ 12 EC Emma Greenwell (10407995)
Biological sciences: Limnology and oceanography Supervisor: Liana Bastos Sales Examiner: Michiel Kraak
20th December 2013 – 27th March 2014
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Table of Contents
I. Abstract .............................................................................................................. 4
II. Introduction ....................................................................................................... 5 2.1 What are phthalates? ................................................................................................................................ 5 2.1.1 Common phthalates ............................................................................................................................... 5
2.2 Environmental fate of phthalates ........................................................................................................ 6 2.2.1 Differences in seasons ............................................................................................................................ 8
2.3 Levels in the environment ...................................................................................................................... 8 2.4 Half-‐lives ......................................................................................................................................................... 9 2.5 Inside the organism ................................................................................................................................ 11 2.6 Modes of action once inside an organism ..................................................................................... 11 2.7 Environmental risk limits .................................................................................................................... 12 2.8 Objective ...................................................................................................................................................... 12 III. Method ........................................................................................................... 13
IV. Results ............................................................................................................ 13 4.1 Summary of literature (1980-‐1999) ............................................................................................... 14 4.2 Literature (1980-‐1999) ........................................................................................................................ 16 4.3 Summary of literature post 2000 ..................................................................................................... 17 4.4 Literature post 2000 .............................................................................................................................. 19 4.4.1 DEHP .......................................................................................................................................................... 19 4.4.2 DBP ............................................................................................................................................................. 24 4.4.3 DEHP and DBP ....................................................................................................................................... 28 4.4.4 DINP and DIDP ...................................................................................................................................... 29
V. Discussion ........................................................................................................ 29 5.2 DEHP ............................................................................................................................................................. 30 5.3 DBP ................................................................................................................................................................ 30 5.4 Nominal concentration experiments with DEHP and DBP .................................................... 31 5.5 DINP and DIDP .......................................................................................................................................... 32 5.6 Exposure routes ....................................................................................................................................... 33 5.7 Problematic variables and environmental risk limits ............................................................. 33
VI. Conclusions ..................................................................................................... 34 6.1 Classification of phthalates .................................................................................................................. 34 6.1.1 DEHP .......................................................................................................................................................... 35 6.1.2 DBP ............................................................................................................................................................. 35 6.1.3 DINP and DIDNP ................................................................................................................................... 35
4.9 Recommendations ................................................................................................................................... 35 VII. Author’s remarks ........................................................................................... 36
VIII. References .................................................................................................... 37
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GLOSSARY Environmental risk limit (ERL) – represent the potential risk of the substance to the ecosystem and are derived using data from ecotoxicology and environmental chemistry. Oocytes – a cell in an ovary, which might undergo meiotic division to form an ovum. Vitellogenin – a protein present in the blood from which the substance of the egg yolk is derived. Planktivores – An organism that feeds on plankton. Glucuronides – any substance produced by linking a glucuronic acid to another substance (via glycosidic bonds). This method (glucorinidation) is used by animals to help excrete toxic substances from the body. Environmental risk assessmen (ERA) – An evaluation of the interactions of agents, human and ecological resources. No observed effect concentration (NOEC) – the highest treatment (test concentration) of a substance that shows no statistical effect compared to a control. Predicted no effect concentration (PNEC) – the concentration below which a specified percentage of species in an ecosystem are expected to be protected. Nominal concentration – The concentration if you all test material added to the test solution dissolved. Effective concentrations (EC50) – the concentration of a substance, which induces a response halfway between the baseline and maximum after a specified exposure time. The number refers to the position within the baseline-‐maximum scale. Gonado-‐somatic index – calculation of the gonad mass as a proportion of the total body mass. Spermatozoa – a sperm cell. Spermatocyte – immature male germ cell which undergoes meiosis developme into a sperm cell. Spermatagonia – any cell of the male gonad that mature to form spermatocytes. Hypertrophy – a non-‐tumorous enlargement of an organ (or part) as a result of increased cell size rather than cell number. Spiggin – a glycoprotein glue used by three-‐spined sticklbacks to stick their nests together. Peroxidation – a chemical reaction in which oxygen atoms are formed leading to production of peroxides. Photodegradtion /photodegradable – substances capable of being chemically broken down by prolonged exposure to light. Octanol-‐water partition coefficient (Kow) – a coefficient representing the ratio of the solubility of a compound in octanol to its solubility in water. Soil organic carbon-‐water partitioning coefficient (Koc) – the ratio of the mass of a chemical that is adsorbed in the soil per unit mass of organic carbon in the sol per the equilibrium chemical concentration in solution. Phytoremediation – the use of plants to remove/neutralize contaminants.
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I. Abstract Phthalates are endocrine disrupting compounds produced on a mass scale for use in plastics. They are not chemically bound to the product and therefore leach into the environment exposing fish to a range of endocrine toxicities. Environmental risk limits (ERLs) are difficult to calculate as different solubility, exposure method, fish species and even age all combine to produce different toxicity effects. In most literature environmental phthalate levels were above the ERL. This paper focuses on what are associated endocrine toxicity effects (metabolic, developmental and reproductive) of di-‐2-‐ethyl-‐hexyl phthalate (DEHP), di-‐butyl phthalate (DBP), di-‐isononyl phthalate (DINP) and di-‐isodecyl phthalate (DIDP). Results consist of 18 studies on phthalate toxicity filtered to only include results from DEHP, DBP, DINP and DIDP on fish species. A mixture of effects on growth inhibition, VTG level alteration, inhibition of oocyte maturation, increased mortality, spinal deformities and maturation inducing hormone alterations etc. were observed with all both DEHP and DBP. Effects were seen to be more potent in pre/early life exposure compared to adults and sometimes even irreversible. Both DEHP and DBP phthalates produces developmental toxicity effects such as increased mortality, retardation in ovary development, decreases in body weight and length, inhibition of 5α-‐adione, decreases in fertility and many more. The order of literature available went DEHP>DBP>DINP/DIDP. For the latter two (DINP/DIDP) only one study was found post year 2000. The availability of DEHP and DBP information allows to derive reasonable ERLs values. However due to the lack of DINP/DIDP information DEHP is used as a proxy for DINP/DIDP ERLs. In conjunction, there is no uniform exposure route to which ERL’s are based on and as seen in the results different exposure routes of the same compound can produce different effects. More solid guidelines of phthalate testing are needed on all compounds especially those of DINP and DIDP.
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II. Introduction 2.1 What are phthalates? Phthalates are chemical compounds used to reduce the chemical affinity between plastic molecules therewith increasing the flexibility of the product sometimes making up 50% of the finished plastic product (Oehlmann et al., 2009; OEHHA, 2009). They are also known to be endocrine disrupting compounds (EDCs) (Ikele, 2011). EDCs may be natural or synthetic compounds that interfere with endocrine regulated processes such as growth and reproduction (Crain et al., 2008). The international program for chemical safety defines endocrine disrupters as “exogenous substances that alter function(s) of the endocrine system and consequently cause adverse health effects in an intact organism or its progeny secondary to changes in the endocrine function” (ECPI, 2009). Production of phthalates consists of around 1 billion tones per year worldwide. They are present in the medical environment, cosmetics, computers, children toys, food packaging, car products and paint making them an unavoidable part of modern life (Mankidy et al., 2013; OEHHA, 2009; Guven and Coban, 2013 and Carnevali et al. 2010). Phthalates are not chemically bound to the plastic molecules within the product meaning they are able to leach out into the environment rendering these compounds unstable within their plastic counterpart (Oehlmann, et al., 2009 and Mankidy et al., 2013). Consequently phthalates are ubiquitous the environmental and ecological concerns surrounding them are increasing.
2.1.1 Common phthalates The general structure of phthalates can be seen in figure 1 (to the right) (R-‐alkyl chain). The most common phthalates are di-‐n-‐butyl phthalate (DBP) and di-‐2-‐ethyl-‐hexyl phthalate (DEHP) (Jarmolowicz et al., 2013; Huang et al., 2008 and Uren-‐Webster et al., 2010). These two specific phthalates
Figure 1: General structure of phthalates (Ogunfowokan et al., 2006)
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occur at higher concentrations than other phthalates (Van Wezel et al., 2000) and have the highest toxicity (out 8 common phthalates under the U.S. environmental protection agency (EPA) management plan) to terrestrial and aquatic organisms (EPA, 2012). These two phthalates produce reproductive and developmental toxicity effects (Jarmolowicz et al., 2013; Lee and Liang 2011 and Zanotelli et al., 2009). Newer phthalate compounds such as di-‐isononyl phthalate (DINP) and di-‐isodecyl phthalate (DIDP) have shown to have no (or very low) toxic effects on aquatic organisms (EPA, 2012; Oehlmann et al., 2009 and Hallmark 2010) despite the reproductive development effects in two generations of rats (OEHHA, 2010).
2.2 Environmental fate of phthalates Once in the environment phthalates are transported through water where they may be dissolved (water sink) or due to its low solubility end up within the sediment (Huang et al., 2008). Here the phthalate compounds are transferred to fish and other aquatic organisms through their diet or by water (Jarmolowicz et al., 2009). Benthic feeders contain higher levels of phthalate compounds within their system compared to planktivores due to the low solubility of most phthalates (Huang et al., 2008; Oehlmann et al., 2009; Mankidy et al., 2013 and OEHHA, 2009). The levels of phthalates within water are affected by water quality such as chemical oxygen demand, dissolved oxygen, ammonia-‐nitrate, suspended solids etc. (Haung et al., 2008). Each phthalate has a different molecular weight that also gives it different properties. A high molecular weight (HMW) means that the compound may be less biologically available while low molecular weight (LMW) compounds are more biologically available (Berge et al., 2013). This makes sense with some literature as DBP (MW 278.4g/mol) has a lower molecular weight then DEHP (390.6g/mol) so therefore is more available for uptake (Teil et al., 2012). In France three fish species were analyzed to see which phthalates were more abundant (Teil et al., 2012). Contradictory to Huang et al., 2008) DBP was the main phthalate found in roach (Rutilus rutilus) followed by
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DEHP. This would confirm the theory that LMW compounds are more readily biologically available than HMW. The gradients for soil was however opposite with DEHP being the main phthalate, but this too would fit theory that phthalates with a low log Kow (inverse of octanol-‐water partition coefficient, related to aqueous solubility) are better at forming solutes (dissolving) than phthalates with a high log Kow. DBP has a log Kow of 4.75 while DEHP has a higher one at 7.5. Phthalates with a high log Kow are more likely to have a higher % in the sediment as the particles that do not dissolve sink towards the sediment within a water column (Berge et al., 2013). As DEHP has a higher log Kow it means that it will be present in larger quantities compared to DBP in sediment samples.
When looking at the log Kow of DINP and DIDP both have a value of 8.8. This value may be derived from another phthalate, which makes it unreliable toward the specific phthalate (ECPI 2014 and Megaloid1 2013). All in all more attention should be placed upon sediment as it tends to have the highest levels, even during different seasons (Figure 5) (Sibali et al., 2013). All phthalates however have a low solubility meaning that once saturated in the water, particles of phthalate will join the sediment (Sibali et al., 2013). Figure 5 shows
Figure 5: Sediment and water levels of phthalates (DEHP, DBP, DEP and DMP at different sample sites along the River Jeksei during two seasons (Sibali et al., 2013).
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the differences in water and sediment phthalate levels from the River Jukskei, South Africa.
2.2.1 Differences in seasons It is still unclear why these differences in seasons arise. For atmospheric phthalates for example seasonal differences can be due to influences of emission sources such as the burning of coal in cold season that would then produce phthalate particulates in the air (Kong et al., 2013). Another reason could be a meteorological parameter. Intense sunlight during the summer, when photochemical reactions are increased and degrade phthalates lowering the concentrations within the atmosphere. Rain can also be a culprit through diluting and washing away phthalates particulates (Kong et al., 2013). When comparing the water and sediment levels in the graph above it is possible that the high winter levels are due to a lack of rain therefore concentrating the phthalates. African summer (rain period) could perhaps dilute the phthalate concentrations within the water and sediment therefore lowering the concentrations (Sibali et al., 2013). Plants have also very recently been shown to significantly enhance the dissipation of phthalates in soil in three ways: phytoremediation, increased sorption of phthalates to soil and plant promoted biodegradation (Li et al., 2004). This could be another explanation for the lower summer concentrations of phthalates in figure 5. Half-‐lives of phthalates can also be increased through increased sorption and cooler temperatures (Staples et al., 1997 and Kickham et al., 2012).
2.3 Levels in the environment In the 1990’s the levels of phthalates in river water, in Manchester, UK for example, were at a mean of 21.5μg/L ±12.5 and 1.3μg/L ±0.9 for DBP and DEHP respectively (Fatoki and Vernon, 1990). High standard deviation was due to the different sample station along the river Irwell. However surprisingly levels at the effluent of a sewage treatment plant were the lowest at 6μg/L for DBP while all other
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sample sites were above 12.1μg/L. For DEHP the highest concentration was found at the sewage treatment plant (1.9μg/L) that also coincided with the percentage of DEHP found in the samples 1.9% for DEHP (79.4% for DBP). This contradicts previous research claiming that DEHP has the highest environmental levels. However this could be due to the higher degradability of DEHP under anaerobic conditions (Huang et al., 2008). In Germany DEHP surface water levels ranged between 0.33-‐97.8μg/L and sediment levels varied between 0.21-‐8.44mg/kg dry weight and for DBP 0.12-‐8.80μg/L and 0.06-‐2.08mg/kg dry weight, respectively (Fromme et al., 2002). This study showed both phthalates to have a wide variability in levels throughout Germany although DEHP always had the highest levels. In the Netherlands environmental measurements were taken in 2005 and it was found that DEHP showed the highest concentrations in both mean municipal sewage treatment plant and industrial waste water types (Vethaak et al., 2005). Mean municipal sewage treatment plant effluent levels were around 1.5μg/L and industrial wastewater levels were 150μg/L compared to DBP that showed levels of 0.3 and 2.2μg/L. 70% of the DEHP and DBP samples contained levels above the level of detection (LOD) although only 30% of the DBP samples were above the LOD in the sewage treatment plant effluent water. Fish muscle concentrations of Bream (Abramis brama) and Flounder (Platichthys flesus) showed mean concentrations of 0.044μg DBP/g, 0.153μg DEHP/g and 0.0078μg DBP/g, 0.064μg DEHP/g in each fish respectively (Vethaak et al., 2005). It seems that phthalate concentration varies not only within country or city but also within micro environments and water types.
2.4 Half-‐lives Half-‐lives of phthalates are the time for a substance to fall to half its original concentration (i.e. degrading) (Staples et al., 21997) through hydrolysis of ester bindings (Liang et al., 2008) and the range of half-‐lives referring to phthalates is vast. Staples et al., (1997) reported a half-‐life of 28 day on average for phthalates within sewage sludge,
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while within the atmosphere half-‐lives consist of around one day (DBP-‐<6 days, DEHP-‐<2 days, DINP-‐<2 days). Within sediment half-‐lives of approximately <one week – several months may be recorded and within surface waters <one day – two weeks (Staples et al., 1997). Staples et al. also reported a half-‐life of years through aqueous hydrolysis (DBP–22 years, DEHP-‐2000 years). In contrast Yuan et al. (2010) reported that DBP and DEHP had half-‐lives of 1.6-‐2.9 days and 5.0-‐8.3 days within sediment, respectively. It has also been postulated that DEHP degrades fairly rapidly under aerobic conditions (Brooke et al., 1991). Microbial degradation has shown DBP to be completely degraded within 28 days (Liang et al., 2008). In Turner and Rawling (2000) eight phthalates were found in a water sample and half-‐lives were measured. On average the phthalate half-‐life in aerobic conditions was between 2.4-‐14.8 days and 14-‐34 days under anaerobic conditions. Other studies such as Yuwatini et al., (2006) showed that DEHP half life in water is approximately two days while in sediment it can last up to 14 days. Magdouli et al., (2013) stated that half-‐lives of DEHP are <one month in aerobic conditions and >one month in anaerobic conditions. In water (with sun) under acidic conditions half-‐lives can be around 390 days while in neutral conditions may be up to 1600 days. From above it is clear that phthalate half-‐lives may have wide ranges. This is due to the different environmental compartments in which the phthalate may be present i.e. atmosphere, sediment, water, inside the organism as each situation will affect the half-‐life as well as what process of degradation is measured. This makes it difficult to consent on fixed half-‐life values. In general it is thought that the longer the phthalate chain (R group in figure 1) the longer the half-‐life and the more persistent it will be and that aerobic conditions will almost most certainly speed up degradation compared to anaerobic (Liang et al., 2008). The organization for economic co-‐operation and development (OECD) has guideline tests and criteria for defining ‘ready biodegradability’. Using these criteria, >60% removal of inorganic carbon within a 10-‐day window of the 28-‐day test, both DBP and DEHP are readily biodegradable in all three states (water, sediment and air). Data concerning DINP was only available for
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atmospheric half-‐life but still fits within the criteria for bing readily biodegradable. If all half-‐life tests incorporated these test guidelines then more accurate comparisons could be made.
2.5 Inside the organism Phthalate accumulation within organisms is also low, partly due to their biodegradability but also due to the compound itself not being highly accumulative in tissue, rendering phthalates non bio-‐accumulative compounds (Van Den Berg et al., 2003; Oehlmann et al., 2009 and Mankidy et al., 2013). Due to their high transformation rate phthalates are not bio-‐accumulative (Mankidy et al., 2013 and Van Den Berg et al. 2003) meaning that on one hand the phthalate compound is transformed into a metabolite that can then interact with receptors and enzymes within the organism (Euling et al., 2013). On the other hand, this metabolism also produces sulphates and other glucuronides that assist in the removal of the parent compound (phthalate) reducing the adverse effects of the phthalate to the organism and also through the food chain (Van Den Berg et al., 2003 and Van Wezel et al. 2000).
2.6 Modes of action once inside an organism Phthalates being EDC’s have a multiple array of modes of action (MOA) making it important to understand how the EDC interacts on a cellular level (Nelson and Habibi, 2013). Endogenous hormones (specifically estrogen and androgen) are most commonly the concern when regarding phthalates. Estrogenic receptors (ERs) and androgenic receptors (ARs) are important in reproduction (ER and AR), sexual differentiation (AR) and even adult sexual behavior (AR) (Harbott et al., 2007 and Thibaut and Porte, 2004). Peroxisome proliferator activated receptors (PPARs) act as regulators for lipid and carbohydrate metabolism as well as cell differentiation (Maradonna et al., 2013). Another MOA is through oxidative damage (OxD) that can cause disturbances to the cellular metabolism (Harbott et al., 2007). All these receptors are present on cell walls. EDC’s show similar biological effects to estrogens and androgens and interfere (agonistically/antagonistically) with the cell receptors (Van
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den Berg et al., 2003) either decreasing or increasing gene expression, production of hormones, enzymes and phase II metabolites affecting the level of active hormones present within an organisms (Thibaut and Porte, 2004).
2.7 Environmental risk limits The European commission previously considered the four phthalates (DEHP, DBP, DINP and DIDP) priority substances meaning that environmental risk assessments (ERA) must have been carried out on these substances (Oehlmann et al., 2008). ERA’s compare environmental concentrations or predicted environmental concentrations (PEC) with the predicted no-‐effect concentrations (PNEC). When the PEC/PNEC ratio is <1 there is no risk, where as if the ratio is ≥1 there is a potential risk meaning strategies must be put in place to reduce the concentrations. For the EPA to recognize acute effects, a total of five tests must be completed on at least four different species using the limit of solubility concentration (max. 3μg/L) (Oelmann et al., 2008). By 2004 the European union risk assessment reports stated that for DBP, DINP and DIDP there was no need for testing or information. DEHP was not granted similar status and therefore still remained on the priority substance list in 2008 (EC, 2014 and Oehlmann et al., 2008).
2.8 Objective This paper will focus on plastic derived EDC known as phthalates. Background on phthalates and why they are the focus of research will be given. It will highlight the associated endocrine disruptions (developmental and reproductive) and will speculate to future work. In previous reviews, fish have never been the sole focus neither experiment set up explained. It has been approximately 13 years since the last review that incorporated over 12 studies (Van Wezel et al., 2000). The paper will focus on four phthalates allowing a more refined and in depth review.
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III. Method For this paper focus was on the compounds DEHP, DBP DINP and DIDP due to their high abundance within the environment (former two) and acclaimed ‘no effects’ of the latter two (Oehlman et al., 2009 and Hallmark, 2010). Searches were be carried out on ’google’ ‘google scholar’ and ‘Web of science’ focusing mainly on publications within the years 2000-‐2014. Searches for DEHP, DBP, DINP, DIDP, effects of DEHP/DBP/DINP/DIDP on aquatic organisms/fish, reproductive/developmental/metabolic effects of phthalates, vitellogenin effects of phthalates, intersex caused by endocrine disrupters, and environmental phthalates are a few of the search terms used. The main duration of research lasted approximately 3 weeks-‐1 month and only full text articles were incorporated within the paper.
IV. Results A total of 46 papers were gathered and divided into sections on organism toxicity (≈18), phthalate levels in the environment (≈5), general information, however nearly all articles had multiple section uses. Due to the majority of organism toxicity publication a further division of pre 2000 and post 2000 research as well as compound groups were added. This was done due to the majority of papers found being post 2000 and to separate ‘recent work’ from ‘previous work’. All publications were given in publication date order (oldest-‐newest). Most experiments within the ecotoxicology field focus on either in vivo or in vitro set-‐ups. The former refers to the whole organism being studied allowing observation of the overall effect of compounds on the organism. The latter refers to using cells in controlled environments (such as petri dishes, assays, etc) where for example assays can provide information on the mechanism of action (MOA) of certain compounds; unfortunately this does not mimic the whole organism (Sohoni and Sumpter, 1998).
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4.1 Summary of literature (1980-‐1999) Table 1: ED effects of phthalates in in vitro receptor binding affinity tests.
Cell type Effect (mM) Remark Original references (within Van Wezel et al., 2000)
DBP Trout hepatocyte EC50 = 1 REP: 6.7x10-‐6 Jobling et al., 1995
Trout hepatic cytosol EC 10-‐25 = 0.17 REP: 2x10-‐5 Knudsen and Pottinger, 1999 DEHP
Trout hepatocyte EC75 = 1 REP: 1x10-‐5 Jobling et al., 1995
Trout hepatic cytosol EC10-‐25 = 0.17 REP: 2x10-‐5 Knudsen and Pottinger, 1999 DINP
Trout hepatic cytosol No effect at 0.17 -‐ Knudsen and Pottinger, 1999
REP: relative potency compared with 17-‐estradiol (Based of appendices by Van Wezel et al., 2000). Table 2: Toxicity data for DBP.
1-‐Y: chemical analyzed in test solution and N: chemical not analyzed in test solution or no data. 2-‐S: static, R: Static with renewal and F: flow through. 3-‐S: survival, R: reproductive and G: Growth. *-‐Average of results (mg/L) when all parameters and authors were the same (Based of appendices by Van Wezel et al., 2000).
Organism Chemical analysis1
Test type2
Exp. time
End point3
Results (mg/L)
Original references (within Van Wezel et al., 2000)
Chronic toxicity to freshwater organisms: NOEC values Oncorhynchus mykiss Y P 60d G 0.1 Rhodes et al., 1995
Pimephales promelas Y F 20d G 0.56 McCarthy and Whitmore, 1985
Acute toxicity to freshwater organisms: L(E)C50 values Brachydanio rerio Y S, R 96h S 2.2 Scholz, 1994 Lepomis macrochirus* N S 96h S 1.6 Mayer and Ellersieck, 1986
Lepomis macrochirus N F 96h S 1.6 Mayer and Ellersieck, 1986 Oncorhynchus mykiss* N S 96h S 4.4 Mayer and Ellersieck, 1986
Oncorhynchus mykiss N F 96h S 1.5 Mayer and Ellersieck, 1986 Oncorhynchus mykiss Y -‐ 96h S 1.2-‐1.8 Hrudey et al., 1976 Oncorhynchus mykiss Y F 96h S 1.6 Adams et al., 1995 Perca flavescens N F 96h S 0.35 Mayer and Ellersieck, 1986 Pimephales promelas N S 96h S 1.3 Mayer and Ellersieck, 1986 Pimephales promelas N F 96h S 4 Mayer and Ellersieck, 1986
Pimephales promelas N -‐ 96h S 2 McCarthy and Whitmore, 1985
Pimephales promelas Y S 96h S 1.5 Adams et al., 1995 Pimephales promelas* Y F 96h S 0.97 DeFoe et al.,1990
Pimephales promelas Y F 96h S 0.92 Adams et al., 1995
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Table 3: Toxicity data for DEHP.
Organism Chemical analysis1
Test type2
Exp. time
End point3
Results (mg/L)
Original references (within Van Wezel et al., 2000)
Chronic toxicity to freshwater organisms: NOEC values Brachydanio rerio N R 35d S, G ≥0.32 Canton et al., 1984 Gasterosteus aculeatus N -‐ 28d S, G ≥0.32 Van den Dikkenberg et al., 1989
Jordanella floridae N S 28d S, G ≥0.32 Adema et al., 1981 Oncorhynchus mykiss Y F 102d S, R 0.005 Mehrle and Mayer, 1976 Oncorhynchus mykiss Y F 90d S, G, R >0.5 DeFoe et al., 1990
Oncorhynchus mykiss Y F 70d S, G, R >0.0073 Cohle and Stratton, 1992 (EU draft)
Oryzias latipes Y F 168d G 0.55 DeFoe et al., 1990 Oryzias latipes N R 28d S, G ≥0.32 Adema et al., 1981 Pimephales promelas Y F 56d S, G 0.062 Mehrle and Mayer, 1976 Poecilia reticulata N -‐ 28d S, G ≥0.32 Adema et al., 1981
Acute toxicity to freshwater organisms: L(E)C50 values Brachydanio rerio N -‐ 96h S >0.32 Van den Dikkenberg et al., 1989 Brachydanio rerio Y R 96h S >100 Scholz, 1995 Gasterosteus aculeatus N -‐ 96h S >0.32 Van den Dikkenberg et al., 1989
Ictalurus punctatus -‐ S 96h S >10 Mayer and Sanders, 1973 Ictalurus punctatus Y F 96h S >100 Johnson and Finley, 1980 Ictalurus punctatus N S 24h S >100 Mayer and Ellersieck, 1986 Ictalurus punctatus N S 96h S >100 Mayer and Ellersieck, 1986 Ictalurus punctatus N F 96h S >0.2 Mayer and Ellersieck, 1986 Jordanella floridae N -‐ 96h S >0.32 Van den Dikkenberg et al., 1989 Lepomis macrochirus -‐ S 96h S >10 Mayer and Sanders, 1973 Lepomis macrochirus N S 96h S >250 Bionomics Inc., 1972 Lepomis macrochirus Y F 96h S >100 Johnson and Finley, 1980 Lepomis macrochirus Y S 96h S >0.2 Adams et al., 1995 Lepomis macrochirus N S 24h S >100 Mayer and Ellersieck, 1986 Lepomis macrochirus N S 96h S >100 Mayer and Ellersieck, 1986 Lepomis macrochirus N F 96h S >0.2 Mayer and Ellersieck, 1986 Oncorhynchus mykiss -‐ S 96h S >10 Mayer and Sanders, 1973 Oncorhynchus mykiss -‐ S 96h S >1000 Silvo, 1974 (EU draft) Oncorhynchus mykiss N S 96h S >540 Hrudey et al., 1976 Oncorhynchus mykiss Y F 96h S >0.32 Adams et al., 1995 Oncorhynchus kisutch N S 24h S >100 Mayer and Ellersieck, 1986 Oncorhynchus kisutch N S 96h S >100 Mayer and Ellersieck, 1986 Oncorhynchus mykiss N S 24h S >100 Mayer and Ellersieck, 1986 Oncorhynchus mykiss N S 96h S >100 Mayer and Ellersieck, 1986 Oncorhynchus mykiss Y F 96h S >20 DeFoe et al., 1990 Oryzias latipes N -‐ 96h S >0.32 Van den Dikkenberg et al., 1989 Oryzias latipes Y F 96h S >0.67 DeFoe et al., 1990 Pimephales promelas -‐ S 96h S >10 Mayer and Sanders, 1973 Pimephales promelas Y F 96h S >0.67 DeFoe et al., 1990
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Pimephales promelas N F 96h S >1 Mayer and Ellersieck, 1986 Pimephales promelas Y F 96h S >0.33 DeFoe et al., 1990 Pimephales promelas Y S 96h S >0.16 Adams et al., 1995
1-‐Y: chemical analyzed in test solution and N: chemical not analyzed in test solution or no data. 2-‐S: static, R: Static with renewal and F: flow through. 3-‐S: survival, R: reproductive and G: Growth. EU draft: (DEHP) (Based of appendices by Van Wezel et al., 2000).
4.2 Literature (1980-‐1999) A meta-‐analysis using journals from 1980-‐1999 was carried out by Van Wezel et al., 2000. Above in table 1, 2 and 3 a summary of these results (no observed effect concentration-‐NOEC, X% effective concentrations-‐ECx, chronic and acute exposure) concerning fish can be found. They found minimal difference between nominal and actual concentrations used in studies concerning DBP. The most sensitive freshwater organism was Oncorhynchus mykiss that showed the lowest chronic NOEC at 0.1mg/L (table 2). Acute toxicity data was more available (see table 5) and Van Wezel et al. reported that ‘no useful test’ regarding soil or sediment was found. When comparing chronic and acute DEHP results it was found that both categories showed no effects in the majority of the studies (even at the highest concentration tested acute: 0.55mg/L and chronic: 1x106mg/L). When effects were recorded and NOEC could be produced the NOEC was above the water solubility of phthalates (3μg/L). With all the data available the authors derived an ERL for the aquatic and sediment environments. For DBP this was done by using the lowest NOEC (0.1mg/L) and applying an assessment factor of 10. For sediment due to lack of data the ERL was derived by multiplying the lowest Koc, partition coefficient between organic carbon in the soil/sediment and water, value (1.2x103L/kg: 12mg/kg). For DEHP, due to no effects observed, the NOEC for the only soil organisms (Rana arvalis – frog) was used 10mg/kg fresh weight and applying a factor of 10. The ERL for soil was then used to derive an ERL for water by combining with the lowest soil/sediment Koc. The derived ERLs for DBP: 10μg/L and 0.7mg/Kg fresh weight and DEHP: 0.19μg/L and 1.0mg/Kg fresh weight. When surface water
17
samples were taken at different location in the Netherlands they found that DBP levels were rarely above the ERL (both water and sediment) derived in this study. For DEHP however unexpected levels 3-‐20 times higher than the derived ERL for water were observed and sediment levels were also much higher than the derived sediment ERL.
4.3 Summary of literature post 2000 Table 4: DBP summary.
N-‐Depicts nominal concentrations. A-‐depicts acute exposure studies. C-‐depicts chronic exposure studies. []-‐concentration causing significant effects. VTG-‐vitellogenin. D/hpf-‐days/hours post fertilization. Dep.-‐depuration (none contaminated water). *-‐<0.05, **<0.01, ***<0.001 significance levels.
Species Age, sex, exp. type and concentration (μg/L) (unless indicated)
Exposure route and duration
Effects Authors
Sander lucioperca
Juvenile (61 dph) (in vivo) 0.125, 0.25, 0.5, 1, 2g/Kg feed
Food 5 weeks
*No effects on female fish, growth rate and survival. *Increases in [DBP] shows decreases in male specimens.
Jarmolowicz et al., (2003)
Danio rerio Adult male (in vivo and in vitro) 500
Water 15 days
Day 7: *increase in surface density of peroxisomes.
Day 15: *increase in both surface density and numerical density of peroxisomes. Increase in activity of acyl-‐CoA oxidase.
Ortiz-‐Zarragoitia and Cajaraville (2005)
Danio rerio
1) Embryos (1-‐2 hpf) 25, 100 2) Adult female 100, 500 (in vivo)
Water 1: 8 weeks 2: 15 days
1) [100] Increase in number and volume of peroxisome density and acyl-‐CoA oxidase enzyme. 2) Mortality of female offspring increased.
Ortiz-‐Zarragoitia et al., (2006)
Gasterosteus aculetaus
Adult male (in vivo) N50, N100 (Measured levels 15, 35 respectively)
Water 22 days
[35*/**] Increase in testosterone and oxidised testosterone. Decrease in spiggin (protein glue).
Aoki et al., (2011)
Cyprinus carpio (in vitro) 100μM, 1mM
Incubated in vitro
[100] Inhibited formation of 5α-‐Adione and synthesis of 5α-‐DHT. [1] Increased synthesis of 17α,20α/βDP
Thibaut and Porte (2004)
Pimephales promelis
1 hpf (in vivo and In vitro) 1000
Water 96 hpf None
Mankidy et al., (2013)
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Table 5: DEHP summary.
N-‐Depicts nominal concentrations. []-‐concentration causing significant effects. Hpf-‐hours post fertilization. Dph-‐days post hatch. Dep.-‐depuration (none contaminated water). *-‐<0.05, **<0.01 significance levels.
Species Age, sex, exp. type and concentration (μg/L) (unless indicated)
Exposure route and duration
Effects Authors
Oryzias laptipes
Adult male (in vivo) N0.1, N0.3, N1μmol
Water 2 weeks None
Shioda and Wakabayashi (2000)
Oryzias laptipes
A) a few days old 10, 50, 100 C) 7 month N1, N10, N50 (in vivo and in vitro)
Water A: 5 days C: 3 months
A) [all] VTG protein not present in males. C) [N10*, N50]* GSI lower in females and [Nall] retardation in ovary (oocyte) development.
Kim et al., (2002)
Oryzias laptipes
1dpf (in vivo) N0.01, N0.1, N1, N10
Water Until hatched
[N0.1, N1] Decreased hatch time. Post 5-‐6 months dep.: [N0.01***, N0.1*, N1***] increased mortality, [N0.01*] altered sex ratio. [N0.1*, N1*, N10**] decrease in male body weight.
Chikae et al., (2004)
Poecilia reticulate
<1 week (in vivo) 0.1, 1, 10
Water 3 months
Day 14: [10] decrease in length and weight. Day 49 and 91: [1, 10] decrease in length and weight (more significant in females [all]**)
Zantonelli et al., (2009)
Danio rerio
6 month female (in vivo and In vitro) 0.02, 0.2, 2, 20, 40
Water 3 weeks
[2] Increase in VTG oocytes and decrease in pre-‐VTG oocytes*. [all] down regulation of LHR and plasma VTG.
Carnevali et al., (2010)
Danio rerio
Mature males (in vivo) 0.5, 50, 5000 mg/kg body weight
Injection on day 1 and 5
[5000***] Decrease in fertilization success. [50*, 5000**/***] decrease in no. of spermatozoa and increase in no. of spermatocytes. [5000*] increase in VTG levels (males should not have VTG). [5000**/*] increase expression of acox1 and ehhadh.
Uren-‐Webster et al., (2010)
Cyprinus carpio
(in vitro) 100μM, 1mM
Incubated in vitro
[100] Inhibited formation of 5α-‐Adione.
Thibaut and Porte (2004)
Pimephales promelis
1 hpf (in vivo and In vitro) 1000
Water 96 hpf
[1000**/*] increase in embryo mortality and increased lipid peroxidation
Mankidy et al., (2013)
(Salmo salar)
4 weeks post hatch (in vivo) N0, N400, N800, N1500mg/kg feed
Food 4 weeks
[DEHP] 3x greater then MEHP (metabolite). 1 week after dep.: DEHP+MEHP levels returned to background levels. 1 month dep.: [N1500*] ovo-‐testis combo
Norman et al., (2007)
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Table 6: DINP and DIDP summary.
Species Age, sex, exp. type and concentration (μg/L) (unless indicated)
Exposure route and duration
Effects Authors
Oryzias laptipes
2 week old larvae (in vivo) 20μg/g (for both DINP and DIDP)
Food F0 till gen. F2-‐42dph
DINP: F0 Embryo development showed decreases in red blood cell pigment*. DINP: F1 survival decreased*
Patyna et al., (2006)
Dph-‐days post hatch. *-‐<0.05 significance levels. Patyna et al. 2006 uses two bioassays for a single effect. Therefore only effects proving significant on both assays are used in this table.
4.4 Literature post 2000
4.4.1 DEHP In 2000 Shioda and wakabayashi studied the effects of DEHP (in vivo) on the number of eggs produced by mating couples and number of successful hatchings in medaka fish (Oryzias latipes). For this experiments groups (one male and two females) with the highest number of fertilized eggs were used. Males were exposed for two weeks to low nominal DEHP concentrations of 0.1, 0.3 and 1μmol/L (through means of water) along side a positive (17 β–estradiol, a natural estrogen) and negative control (tap water). Once exposed the males were placed back into their original group. The DEHP concentrations showed no significant effects on number of eggs and hatchings, which could be due to the extremely low concentrations used. Both chronic and acute exposures of DEHP (in vivo and in vitro) were studied by Kim et al. in 2002. Japanese Medaka fish (seven months old) were exposed via water to concentration of 10, 50 and 100μg/L of DEHP (for acute testing). For chronic testing fish a couple days old were exposed to nominal concentrations of 1, 10 and 50μg/L. In acute exposure (5 days) it was found that the protein (200-‐kDa) used for identification of vitellogenin (VTG) proteins were not present in male Medaka in all four exposures (including the control). In females however VTG was found in all the control and the exposed fish, although two out of the five fish in 1μg/L exposed tank showed
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extremely low levels. Overall, acute effects of DEHP on VTG were not significant. The chronic exposure (three months) to DEHP showed the 200-‐kDa protein not to be present in male fish. In females fish however the protein occurred less frequently as DEHP concentration increased. The weight and length of fish used in the chronic exposure showed no statistical difference in all treatments showing DEHP to have no effect on growth. The Gonado-‐somatic index (GSI) of females in both 10 and 50μg/L DEHP treatments was statistically lower than that of the control females while no effect was found on male fish showing DEHP to inhibit the development of Medaka fish ovaries. Histology of both the gonads and ovaries from the chronically exposed fish were also looked at. Here gonads of the male fish were not deformed compared to the control, while the oocytes within the ovaries of female fish were. In the control females, oocytes were developed to either stage two or three (stage three allowing them to be fertilized). In all 1, 10 and 50μg/L DEHP treatments only 37%, 0% and 22%, respectively, of the fish had matured oocytes at stage three compared to 54% of the control – taking note that 10μg/L showed no stage three development. Along side, only 26%, 25% and 12% of the female fish (respectively of 1, 10, 50μg/L DEHP) could reach stage one compared to the control where oocytes development was not stopped (figure 2). This shows the retardation effects in ovary growth of DEHP using environmentally relevant concentrations. In 2004, Chikae et al. also conducted an in vivo study on the negative (irreversible) effects that DEHP exposure using pre-‐hatched Medaka would have on adulthood (5-‐6 months post hatch). Treatments of water containing nominal DEHP concentrations of 0.01, 0.1, 1, 10
Figure 2: Ovaries of females medaka after 3 months. A) control, developed to stage 3 B) DEHP (10μg/L) stuck in stage 1 (Kim et al., 2002)
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μg/L and a control were used to expose 1-‐day-‐old fertilized eggs. Once hatched the fish were transferred to DEHP free water for 5-‐6 months. At the beginning (pre-‐hatch) over 90% of the eggs in each treatment showed signs of eye development (eyeing) except at 10μg/L were only 83% were found eyeing. Of those eggs that had successful eyeing, over 90% continued to hatch in each treatment. The only significant difference was a decrease in hatching time seen at the 0.1μg/L (P<0.005) and 1μg/L DEHP treatments compared to the control. In adulthood, after no DEHP exposure for 5-‐6 months, irreversible effects were significant compared to the control. Post-‐hatch mortality was significantly increased in the 0.01, 0.1 and 1μg/L treatments (P<0.001, <0.05 and <0.001, respectively). Sex ratio within the 0.01μg/L treatment was significantly altered (4m:16f), which may have been due to increased male mortality or feminization. Body weight was significant different in male fish within the treatment 0.1, 1μg/L (P<0.05) and 10μg/L (P<0.01). This study shows the irreversible effects of phthalate exposure in embryonic states of medaka fish. Norman et al., (2007) studied DEHP (in vivo) on Atlantic salmon
(Salmo salar) with nominal concentrations of 0, 400, 800 and 1500mg DEHP/kg feed. Here levels of DEHP and its metabolite mono-‐ethylhexyl phthalate (MEHP) within fish tissue were studied after acute exposure (four weeks) of DEHP. Along side, histological, growth and liver effects were analyzed after one month of depuration (no exposure to DEHP). The DEHP concentration in the fish tissue post acute phase was three times higher than the concentration of MEHP. Control fish that were not
Figure 3: guppy fish at day 49 with treatments above. Grid is 1mm (Zanotelli et al., 2009).
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exposed to dietary DEHP showed low background levels of DEHP (0.016 mg/kg fish) and MEHP (0.020 mg/kg fish). DEHP and MEHP concentrations increased in tissue as treatment concentration increased. Both were eliminated to near background levels one week after the depuration phase. Mortality in all groups was low (4%) and no difference in weight and sex ratio was recorded between the different exposure concentrations. Within each treatment a few fish (1% of 400 and 1500mg DEHP/kg food) were observed anatomically to be slightly different (increased testes size). The only statistically difference recorded was in the treatment group of 1500mg DEHP/kg feed where 6 out of the 202 fish had ovo-‐testis (P<0.014). This study showed that DEHP had no short-‐term effects. Zanotelli et al., (2009) conducted a study focusing on the growth (weight and length) of <1-‐week-‐old (larval) guppy fish (Poecilia reticulata). The guppy fish were subjected to continuous exposure (in vivo) to DEHP through water (0.1, 1, 10μg/L). By day 14 a statistically significant growth inhibition at the highest DEHP concentration was observed and increased with time. After 49 days of exposure, DEHP treated fish were compared to control fish. Length showed a dose-‐dependent decrease, where DEHP exposed fish at 1 and 10μg/L were 15-‐30% shorter (respectively) than the control and weight was decreased by as much as 40-‐70% respectively. After 91 days of chronic exposure to environmentally relevant DEHP concentrations the fish showed a significant decrease in weight and length: fish exposed to 1 and 10μg/L decreased 10% and 26% in length and 32 and 61% in weight, respectively (figure 3 below). There was a higher level of significance within females at day 49, with all concentrations showing a P<0.01, where as with male fish at day 49 only 10μg DEHP/L differed form the control with P<0.01. This study shows that chronic exposure as low as 1μg DEHP/L show a time and dose dependent relationship when it comes to growth. The fish used in this study were considerably small which could have increased the effects observed. Carnevali et al., (2010) experimented on the effects of DEHP using six-‐month-‐old female zebrafish (Danio rerio) in an in vivo and in vitro
23
study. Environmentally relevant concentrations of 0.02, 0.2, 2, 20, 40μg/L as well as a positive control were used to study the impact on fecundity, ovulation and oocytes maturation. Fish were exposed through water to DEHP for three weeks and were compared to a solvent control. Results showed that fish exposed to 2μg/L had a significant increase in the number of vitellogenic oocytes. This was associated with the significant decrease in pre-‐vitellogenic oocytes compared to the control (P<0.05). Down regulation of ovarian luteinizing hormone receptor (LHR) and plasma VTG were significantly different compared to the control at all five doses of DEHP. These two factors clearly show the estrogenic activity of DEHP with regards to the inhibition of oocytes maturation. This is also supported by the dose dependent increase of BMP15, a protein involved in oocytes maturation. After the three-‐week exposure period the female fish were placed into a mating tank with control males, showing that the fecundity of embryos was severely compromised compared to the control. This study shows the concrete risk associated with aquatic organisms living in phthalate-‐polluted areas. Another in vivo and in vitro experiment on DEHP by Uren-‐Webster et al., (2010) studied the reproductive health of male zebra fish. 16 colonies (male and female pairs) were used that were consistent with egg production and spawning were over a 10 day period. Here instead of the dietary or water exposure as previous studies applied, the DEHP solution was injected into the intraperitoneal cavity. This method of administration allowed all fish to receive the same dose as well as being able to target male specimens. Environmentally relevant concentrations of 0.5mg DEHP/kg of body weight (bw), range within measured concentration of wild fish, 50mg/kg bw and an extremely high 5000mg/kg bw was used to assess the mechanisms of phthalate toxicity. All three treatments were compared to a control. The fertilization success of males subjected to 5000mg/kg bw were significantly lower than the other three treatments (P<0.001), although this was only when including the full 10 day exposure period (the first 5 day period showed no significant difference). No abnormal embryo development or embryo survival
24
effects were seen in the treatments. Histological analysis of the gonads showed significantly lower numbers of spermatozoa (sperm cell) in the testes of males injected with 50mg/kg of bw (P<0.05) and 5000mg/kg bw (P<0.01) compared to the control fish. On the other hand there was a significant increase in the number of spermatocytes (immature male germ cell) compared to the control in both 50mg/kg bw (P<0.05) and 5000mg/kg bw (P<0.001). When studying at the liver, a statistically significant increase (P<0.05) in VTG levels was recorded in the treatment 5000mg/kg, which showed DEHP to have estrogenic activity, as VTG should not be found in male zebra fish. In the male fish a significant increase in the expression of the genes acox1 (acyl-‐coenzyme A oxidase 1) and ehhadh (enoyl-‐coenzyme A hydratase/3-‐hydroxyacyl coenzyme A dehydrogenase) that are both involved in lipid metabolism was found. Males showed no alterations in swimming and feeding behavior throughout the study (compared to controls). This study used mature fish which are known to be less sensitive than juvenile fish, which may have caused the conclusion that DEHP at environmentally relevant concentrations (0.5mg DEHP/kg bd) show no short term reproductive effect. Lee and Liang (2011) studied zebra fish offspring and exposed them for 3 months to low doses of DEHP through water in vivo. 2ml of DEHP was placed into tanks containing 110 liters of water, and every month an additional 0.1ml of DEHP was added. They observed that DEHP altered the sex ratio from 1:1 to 3:7, although they failed to specify if this was significant. Decreases in growth (length and weight) were observed, but were however not significant. They concluded that DEHP showed no effect.
4.4.2 DBP In Jarmolowicz et al., (2003) DBP concentrations of 0.125, 0.25, 0.5, 1 and 2g/Kg feed were used to determine the impact on the reproductive system in juvenile European pikeperch (Sander lucioperca) in an in vivo study. A total of 40 fish were placed into each concentration tank with a control tank with no addition of DBP. The
25
experiment was divided over two five week periods the first being 61-‐96 days post hatch and the second 97-‐132 days post hatch. In the first period fish were fed the DBP contaminated feed. During the second period fish were fed uncontaminated feed. 15 fish from each tank were taken for histological analysis at the beginning (60 days post hatch), after the 1st and the 2nd period. There were no negative changes within female fish, nor in survival and growth rates (P<0.05). After 96 days post-‐hatch the sex ratio in treatment groups 0.125 and 025g/Kg feed was 1:1. 50% of the males in those two groups showed gonads that were comparable to those of the control group. The remaining 50% showed smaller testes size, reduced spermatogonia (any cell of the gonad which matured form a spermatocytes) and seminal vesicles. Increasing concentration of DBP showed a positive correlation with reduction in male specimens (P<0.05). Fish within the treatment group 2g/Kg of feed had a significantly altered sex ratio (P<0.05). In the two highest DBP concentration tanks (1 and 2g/Kg of feed) intersex specimens (6.7%) were recorded although not significant. Jarmolowicz et al. concluded that DBP acts as an anti-‐androgen (blocking endogenous androgen action) creating an ‘estrogenic environment’. This study is the first to report DBP disruption in sex differentiation in fish. Ortiz-‐Zarragoitia and Cajaraville (2005) used high DBP concentrations of 500μg/L to observe effects on the liver peroxisomes, enzyme activity of Acyl-‐CoA oxidase and on VTG levels (In vivo and in vitro). They exposed adult male zebra fish through water for 15 days. They found that at day seven the surface density of liver peroxisomes had significantly increased (P<0.05) compared to the control while at day 15 both surface density and numerical density had significantly increased from the control (P<0.05). Acyl-‐CoA oxidase showed a significant increase in activity at both time points (days 7 and 15). Surprisingly DBP showed no significant effect on VTG levels. They concluded that DBP shows no estrogenic effect in male zebra fish.
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The next year (2006) Ortiz-‐Zarragoitia et al., conducted another study (in vivo) on DBP and the Actyl-‐CoA oxidase enzyme, peroxisomes and VTG, but also mortality. This study was conducted in two parts, the first focusing on early life exposure and the second focusing on adult life exposure and their offspring. For the first experiment zebra fish eggs were exposed (via water) to concentrations of 25 and 100μg/L. A solvent control was used to compare results. 1-‐2 hpf eggs were exposed for three weeks. Once hatched they were transferred to a larger tank and exposed for a further five weeks. Measurements were taken at 4, 6, 10 days post fertilization (dpf) and 3 and 5 weeks post fertilization (wpf). Results showed that survival of exposed fish did not differ from the controls. However anatomical deformities were observed in both DBP exposed groups (figure 4). Spinal cord malformations and hypertrophy of the yolk sack were noticed in infant fish and in juvenile fish spinal cord and swim bladder malformations were apparent. Although Ortiz-‐Zarragoitia et al., (2006) fail to specify numbers of malformed fish, however those in the control showed no signs of malformation. As with the prior study in 2005, here too they found that the number and volume of peroxisome density as well as the Acyl-‐CoA oxidase enzyme increased significantly in the 100μg/L treatment at five weeks compared to the control, while no significant differences were recorded in the 25μg/L treatment. All fish within the 25μg/L were male (testes all containing spermatozoa and spermatogenic cells) while only two in the 100μg/L showed both pre-‐vitellogenic and vitellogenic oocytes therefore classified as female compared to the control (6 female and 4 male). Only the 100μg/L treatment caused effects to the fish.
Figure 4: zebra fish A) control at 7 dpf, B) DBP (100μg/L) 7 dpf (Ortiz –Zarragoitia et al., 2006)
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In the second experiment 10 adult female zebra fish were exposed via water for 15 days to 100 and 500μg/L of DBP. After 15 days of exposure each female was paired with two males in untreated water and left to reproduce for two to three days. After spawning the female fish were sacrificed and liver, brain and ovary analysis. Embryos produced during spawning were gathered and placed into the same (treatment) groups as their female parent and then transferred to untreated water for 27 days. The number of eggs produced by the treated females did not differ from the numbers of the control. However, mortality showed a significant dose dependent relationship such as in the highest treatment where 70% mortality was recorded after 25 days. VTG expression, liver VTG protein levels, oocytes and ovary development showed no significant difference compared to the control. Both experiments incorporated mortality of young zerbra fish, however exposure to phthalates pre fertilization increased the mortality where as exposure post hatch showed no affect on mortality. Aoki et al., (2011) conducted the most recent in vivo study on DBP. They chose adult male three-‐spined stickle back (Gasterosteus aculetaus). Fish were exposed through water for 22 days to nominal concentrations of 50 and 100μg DBP/L. Throughout the experiment the concentrations of DBP were measured every three to four days (water samples ran through gas chromatography and mass spectroscopy) where it was found that the actual concentration was much lower than their original calculated input. Mean concentrations of 15 and 35 μg/L were recorded at the 50 and 100μg/L tanks, respectively. There was no significant difference in weight, length or gonado-‐somatic index for either treatment group compared to the control. They did find that testosterone levels and oxidised testosterone levels were significantly higher in the 35μg/L treatment group (P<0.05) compared to the control. Spiggin (protein glue) was also measured in the kidneys, where it was found to have a negative correlation with DBP concentration with only the highest DBP concentration showing a significant decrease in spiggin (P<0.011). A slight delay in nest building behavior of those fish in the 35μg/L
28
treatment group was also found, although this was not significant (three out of eight males failed to build a nest until the last day, 22). Here, high levels of DBP showed increases in testosterone levels, which other studies failed to report, and also a decrease in protein glue.
4.4.3 DEHP and DBP Thibaut and Porte (2004) conducted an in vitro study on steroid synthesis and metabolism using carp (Cyprinus carpio). Here the effects of DBP and DEHP on the enzyme 5α-‐Reductase (5α-‐Re), the maturation-‐inducing hormones (MIH) 17α,20α/β-‐dihydroxy-‐4-‐pregnen-‐3-‐one (17α,20α/βDP), steroid 5α-‐Androstanedione (5α-‐Adione) and the androgenic hormone 5α-‐Dihydrotestosterone (5α-‐DHT) were studied. It was found that both DBP and DEHP (100μM) significantly inhibited the formation of 5α-‐Adione by 45% and 65%, respectively. DBP also significantly inhibited the synthesis of 5α-‐DHT by 41%, while DEHP showed no significant effect. When looking at MIH’s DBP (1mM) significantly increased the synthesis of 17α,20α/βDP, which plays an important role in the oocyte maturation and helps indicate spawning readiness, by around 138-‐220%. Thibaut and Porte concluded that phthalates interfere with the enzymes used to make 17α,20α/βDP, although the most affected pathway was that of 5α-‐Re, therefore decreasing both the steroid 5α-‐Adione and the androgen 5α-‐DHT. This causes alterations in the androgen synthesis, metabolism and male sexual maturation and development. Along side the above, interactions with enzymes used in the formation of 17α,20α/βDP will cause alterations in gamete quality and quantity as well as disrupt the synchronization of spawning and mating behavior (Thibaut and Porte, 2004). DBP shows to be more potent than DEHP in altering sex steroid associated with the reproductive system. Mankidy et al., (2013) used both DBP and DEHP to study embryo mortality and cytotoxicity in 1 hour old fertilized Fathead Minnows embryos (Pimephales promelis) using water concentrations of 1mg/L for both phthalates in this in vitro and vivo study. DEHP at 1mg/L showed a 30% increase in embryo mortality (P<0.01) compared to
29
the control at 3% (P<0.05). DBP did not produce any significant mortality compared to the control. 1 mg/L DEHP also caused a 2-‐fold greater lipid peroxidation than the control (P<0.05), while DBP levels stayed the same as that of the control. DEHP proved to be moderately toxic to Fathead Minnow embryos while DBP showed no significant effects unlike Ortiz-‐Zarragoitia et al., (2006) and Aoki et al., (2011).
4.4.4 DINP and DIDP In 2006 Patyna et al., conduced a multigenerational in vivo (F0, F1 and F2) study on medaka fish using nominal concentration of 20 μg/g DIDP and DINP. Many parameters were measured such as fecundity, growth, embryonic stage development and gonad and hepatic-‐somatic indices. Due to the low water solubility of DINP and DIDP the exposure was through diet, as this would be the likely route in the environment. Long-‐term survival was significantly decreased (P≤0.05) in the F1 generation (DINP). A significant change was the decrease in red blood cell (RBC) pigmentation. The F0 acetone control in both assays1+2 as well as DINP in assay2 showed a significant (P<0.05) delay in RBC pigmentation compared to the untreated control. Within the F1 generation both DINP and DIDP treated fish showed a significant delay in RBD pigmentation (P<0.05) in assay2 compared to the control. However all embryos continued to develop normally. Post hatch larval survival was only significantly affected in one of the F0 DINP treated assays. Body weight, gonad weight, gonadal-‐somatic index and egg production showed no significant differences among the different treatment groups. Overall DINP and DIDP show no chronic effects
V. Discussion Comparing the above studies is difficult, as studies may have administered the phthalate concentration either via food, water or injection as with Uren-‐Webster et al., (2010) and in the case of Van Wezel et al., (2000) exposure routes were not reported. Phthalates with low solubility are more concentrated in the sediment (Berge et al., 2013); therefore benthic dwellers are at a higher exposure and
30
may be subject to increased toxicity. Therefore comparing both diet and water would not be the most effective method for this discussion. Therefore comparisons of each method of exposure (diet, water and injection) will be compared against its own type.
5.2 DEHP When comparing acute DEHP toxicity through water exposure varying results were observed, Kim et al., (2002) concluded no effect from DEHP levels on VTG levels in Medaka while in contrast Carnevali et al., (2010) concluded that there was a significant increase in the number of VTG oocytes as well as a down regulation in the plasma VTG levels. In the case of Uren-‐Webster et al., (2010), where exposure was administered through injection, only the highest treatment group showed significant effects. Zanotelli et al., (2009) showed that unlike the conclusions of Uren-‐Webster, growth was significantly inhibited, showing a dose dependent decrease, where females were more affected than males. Van Wezel et al., (2000) concluded that DEHP had no effect during both chronic and acute exposure. However all studies post 2000, except two (Uren-‐Webster et al., 2010 and Norman et al., 2007) regarding DEHP used concentration under 500μg/L. Yet those studies under 500μg/L showed numerous effects. DEHP shows varying effects with similar and non-‐similar concentrations. Future studies using lower concentrations would benefit in finding the lowest concentration to which certain effects are seen.
5.3 DBP DBP increased liver peroxisome significantly in male adult zebra fish (Ortiz-‐Zarragoitia et al., 2005 and 2006, respectively). In 2006 Ortiz-‐Zarragoitia et al. also stated a concentration dependent mortality in young female zebra fish when exposure was aimed at the pervious female generation. However, when exposed in embryonic state (eggs) there was no difference in mortality. This could be due to the different concentrations used in both parts of the study, as the egg exposure treatments were lower than that of the adults exposure.
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Unlike DEHP, DBP did not affect VTG levels according to Ortiz-‐Zarragoitia et al., 2005 and 2006. In the case of Aoki et al., (2011) nominal concentrations were measured and found to be actually much lower than originally calculated. These concentrations showed no effects on weight, length and GSI however increased levels of testosterone. Dietary exposure showed no effect on female fish, however there were increased reductions in recognized male specimens in both highest concentration treatments. Three of the post 2000 studies (Uren-‐Webster et al., 2005 and 2006 and Mankidy et al., 2013) used concentrations above the NOEC stated by Van Wezel et al., (2000). Like DEHP, DBP varies in its effects with similar and non-‐similar concentrations although DBP shows no effect towards VTG.
5.4 Nominal concentration experiments with DEHP and DBP Experiments that used nominal concentrations not easily compared, as the actual concentration remains unknown. In the case of Aoki et al., (2011) their calculated nominal concentrations showed to be 35% lower than organically calculated. This shows that nominal concentrations may not be accurate. This shows the important of measuring actual phthalate concentrations used or even better the internal phthalate concentrations. However although no accurate concentrations are given, effects were still observed and therefore should be taken into account. Effects such as a decrease in hatch-‐time, increase in mortality, changes in sex ratio, and changes in weight were all part of the irreversible effects of DEHP (Chikae et al., 2004) after 5-‐6 months depuration. The nominal concentrations in this experiment were very low and if inferring percentage changes from Aoki et al., (2011) concentrations would be even lower. In contrast, results reported by Shioda and Wakabayashi (2000) showed no effect in hatching time at the same nominal concentrations. Of course age of the fish was different in both studies as well as gender. In the case of diet exposure (Norman et al., 2007), they too added a depuration phase to
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their study and found that levels of DEHP dropped back to background levels within one week. This contradicts Chikae et al.’s conclusion where there was no sign of increase mortality. Unlike Chikae et al. (2004) and Zanotelli et al. (2009) there were no effects on the growth (length and weight) in the study by Kim et al. 2002. However the GSI was significantly lower in females. The nominal concentrations of DEHP and DBP by Thibaut and Prote (2004) affected the maturation–inducing hormone as well as the levels of enzymes used to make that hormone. These alterations lead to changes in metabolism, development and male sexual maturation (quality and quantity reduction in gametes). Both phthalates (DBP and DEHP) inhibited the formation of the steroid (5α-‐Adione), while only DBP affected the androgenic hormone 5α-‐DHT and increased the synthesis of maturation inducing hormone (17α,20α/βDP). In Mankidy et al. (2013) DEHP showed a significant decrease in mortality and increase in lipid peroxidation, while DBP showed no effect. Perhaps the effects measured were only targeted by DBP. These two shows study shows the different targets of each phthalate. Like both measured DBP and DEHP concentration studies, those that used nominal concentrations found varying effects
5.5 DINP and DIDP In the case of DINP and DIDP there is hardly any literature to compare. Much research is conducted in the human field (Kruger et al., 2012; Kransler et al., 2013 and Silva et al., 2013), rats (Clewell et al., 2013) and even bacteria (Park et al., 2009) but hardly any with regards to aquatic organisms. The only study found was that of Patyna et al. (2006) where two assays were used per treatment, therefore only significant results on both assays will be used in the discussion. The F0 generation showed a significant RBC pigmentation delay although all embryos continued to develop normally. In this study no endocrine mediated effects (testis-‐ova, intersex or sex reversal etc.) were observed, similar to the meta analysis by Van Wezel et al. (2000) showing that DINP and DIDP are not expected to produce chronic effects. Due to the low solubility of DINP and DIDP
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(Kow 8.8) it could be that these phthalates are more available within sediment rather than dissolved in water which would explain the lack of effects through water exposure
5.6 Exposure routes Comparing results obtained based on the same exposure routes still showed no similarity in outcome of the different studies. As mentioned in Uren-‐Webster et al. (2010) it was claimed that use of injections, as an exposure route would be a more accurate administration method to assure all specimens received the same dose. Diet exposure does not guarantee equal dosing, therefore affecting the observations and reproductive or developmental or metabolic measurements of experiments. Water exposure may carry similar disadvantages as diet exposure, as it may be possible that different species of fish consume, ingest or absorb larger amount of water that would affect the level of phthalate present within the organism (Aoki et al., 2003).
5.7 Problematic variables and environmental risk limits Along side nominal concentrations being an inaccurate measure of phthalate exposure concentration, there are also other ways phthalate concentrations can change throughout an experiment. Photodegradation (ability to be chemically broken down by light) and microbial degradation as well as loss of phthalates through adsorption to glass tanks (Staples, 2003) are just two ways contributing to phthalate losses can be ‘lost’ (Kim et al., 2002). Another problem arises when comparing different studies, as different species of course react to contaminants in different ways (Law et al., 2006 and Veltman et al., 2005), which could explain the differences between the studies. Reported here age can also be of importance in determining the severity of toxicity. Young and embryonic exposure seem to give more potent results than exposure in adult life (Zanotelli et al., 2009; Ortiz-‐Zarragoitia et al., 2006 and Chikae et al., 2004).
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None of the studies reported here were dealing with the exposure route of diet by food chain (smaller organisms). This exposure route could perhaps produce different results, where the smaller organism may not be affected as much as the predator that feeds up on the organism and subsequent biomagnification phthalate concentrations (Teil et al., 2012). Exposure route and species are thus important. It must also be taken into consideration that the parent compounds themselves are not the sole cause of toxicity, as their metabolites are quickly transformed within the organism (Euling et al., 2013). All these variables make it hard to produce ERL for the environment. In Van Wezel et al. (2000) it was shown that concentrations above the ERL (3μg/L) showed no effect, however it is not known whether sediment dwelling species were chosen. Concentrations above the water solubility for phthalates would increase the particulate phthalate concentration therefore increasing phthalate levels in the sediment.
VI. Conclusions Due to the hard separation between reproductive and developmental effects, definitions will be used to categorize each effect. Reproductive effects will be defined as any effect to alter sexual function and fertility. Developmental effects will be any that cause structural or functional alterations that may affect growth and differentiation. These definitions are interlinked, which should be taken into account as developmental toxicity may cause reproductive toxicity. From the summary of results (table 1) one can see there are many effects associated with phthalates, especially DEHP.
6.1 Classification of phthalates When it comes to what classification the studies give to the different phthalates it obviously varies. Uber-‐Wester et al. (2010) and Carnevali et al. (2010) both claim DEHP to be an estrogenic compound while DBP, according to Jarmolowicz et al., 2013 and Gray et al., 2000 is considered to be an anti-‐androgen. Single classifications for phthalates are nearly impossible as they nearly always have multiple pathways (ER, AR, PARR, oxidative damage
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etc.) making it extremely hard to classify them into androgenic, estrogenic, anti-‐androgenic etc. compounds. This is because the cell receptors, ERs for example can be regulated by estrogenic and androgenic compounds (Nelson and Habibi 2013).
6.1.1 DEHP Using the definitions above, all but two effects are placed within the developmental category. Decrease in fertilization success remains in reproductive toxicity. Therefore DEHP causes developmental toxicity effects.
6.1.2 DBP DBP, like DEHP, also shows to have the majority of its effects in the developmental category according to the definition used. The only effect within the reproductive realm is the increase in offspring mortality.
6.1.3 DINP and DIDNP DIDP showed to have no effects to fish. DINP showed to cause delayed in the red blood cell pigment and an increase in mortality showing both reproductive and developmental effects. Whether the latter is caused by the former remains to be tested.
4.9 Recommendations Through this research a lack of metabolic studies were found, perhaps search terms used were not specific or simply the research regarding the metabolic effects has not been studied as much as others. This paper only focused on four phthalate compounds although there are over 18 commercial phthalates (Peijnenburg and Struijs, (2006) which could all have differing effects such as DEP which causes increased liver mass along with destruction of kidneys and liver tissue (Iekel, 2011). Most research found was on DEHP while other compounds lacked the same amount of attention. More research should be conducted on DBP and especially on DIDP and DINP, as only one article was found concerning DINP and DIDP effects. From this paper it is clear that perhaps the ERAs need to have a clear protocol that can be easily replicated. This would allow studies an easier comparability and would also allow values concerning NOECs ECs and ERL can be more accurately calculated. Research for this paper was only carried out for one month, given
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more time an increased number of articles would have provided larger consensus view on the topic. The organization REACH (registration, evaluation, authorization and restriction of chemicals) came into force in 2007 and has helped to improve the legislative framework of chemicals in the EU. Since its initiative DEHP containing toys and objects have been restricted in many countries (REACH, 2011). Restrictions such as rejection of import by customs, voluntary stop of sales and voluntary withdrawal of product are just a few examples. Countries that stand out as having more products restricted are Spain, Germany and Finland. The majority of DEHP products have been within a restriction list since 2011 (Reach, 2011). For DBP restrictions have even gone as far to allow voluntary destruction of product by importer to a set of plastic animal toys since 2010 (Reach, 2011). DINP and DIDP containing products have few placements on the list.
VII. Author’s remarks Overall it would seem that DEHP and DBP produce acute and chronic developmental effects such as increased mortality in both adult and offspring, intersex, increase in number and volume of peroxisomes, GSI decreases in females, retardation in ovary development, decrease in growth (weight and length) and decreases in spermatagonia are only just a few. However age of fish, species of fish, exposure route, and diet all play an important role as to how intense and potent the phthalate concentration are. More studies and test are needed in the future to fully understand the complicated mechanism and disruptive effects of each phthalate. Standardized tests would allow all phthalate-‐orientated experiments and measurement to be comparable. This would produce more accurate, narrow and pin pointed results to half-‐lives, endocrine effects etc. to what are already highly endocrine disruptive compounds.
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