Toxic Effects of Ammonia, Nitrite, and Nitrate to Decapod Crustaceans: A Review on Factors Influencing their Toxicity, Physiological Consequences, and Coping Mechanisms.
Nitrogenous wastes including ammonia-N, nitrite-N, and nitrate-N are increasingly becoming a global issue in aquatic ecosystems due to escalating anthropogenic activities and are a ubiquitous concern in aquaculture. These pollutants are interrelated via the nitrification cycle, with the direct metabolic product ammonia-N generally being the most toxic with high species specificity. Furthermore, while environmental factors influencing nitrogenous waste toxicity are similar, the causative underlying mechanisms are often substantially different. In this review, we focus on decapod crustaceans due to their high commercial value and likelihood of encountering these pollutants in their benthic or near-benthic habitat. While a large body of publications exists in this area, to date a comprehensive literature review on relative toxicities of all three nitrogenous wastes, physiological consequences, and adaptive mechanisms of crustaceans is lacking. Understanding these processes will likely have implications for environmental/fisheries management and the aquaculture industry. Additionally, there are strong indications that theoretical "safe" values, traditionally used for predicting toxicity thresholds, substantially underestimate the impact of nitrogenous waste on the growth and physiological condition of crustaceans. These consequences will be emphasized along with various methods of uptake, elimination, and detoxification that ultimately explain differences in nitrogenous waste toxicity to decapod crustaceans.
Keywords: crabs, shrimp; osmoregulation; gas exchange; immunology; histology
Ammonia, nitrite, and nitrate are essential life sustaining components to various aquatic microorganisms and within aquatic ecosystems are the most common dissolved inorganic nitrogen ions. However, excessive levels can significantly affect the abundance and physiological condition of various aquatic animals, including decapod crustaceans, which is becoming a larger issue with increasing global anthropogenic activities (Camargo and Alonso, ). Moreover, these compounds are a ubiquitous concern and major limiting factor in the crustacean aquaculture industry, particularly since the trend in aquaculture is a move toward more intensive systems with a reliance on higher feed inputs (Bouwman et al., 2011; FAO, ).
The origin of these nitrogenous compounds are mostly from the natural decomposition of organic matter and metabolic by-products of protein catabolism yielding ammonia-N ("ammonia-N" referring to both NH4+ and NH3; Camargo and Alonso, ). In aquatic environments, aerobic nitrification naturally converts ammonia-N to nitrite-N (NO2−) and then to nitrate-N (NO3−), as follows:
Generally, of the three, ammonia-N is the most toxic form to decapod crustaceans, followed by nitrite-N and nitrate-N in descending order (Chen and Lei, ; Chen et al., ,; Meade and Watts, ; Romano and Zeng, ). In addition to impacting survival or growth, these pollutants can disrupt various physiological mechanisms in similar ways although the underlying causes can be substantially different depending on the nitrogenous waste species (Jensen, ; Cheng et al., ). Furthermore, among these nutrients, the modes of uptake and elimination vary as well as these processes being influenced by both biotic (e.g., developmental stage and osmoregulatory abilities) and abiotic factors (e.g., temperature, pH, and salinity; Jensen, ; Weihrauch et al., ). This is particularly important since the ability of crustaceans to resist uptake or effectively eliminate these pollutants from their hemolymph, will in turn greatly influence both their physiological condition and tolerance to nitrogenous waste exposure (Jensen, ; Weihrauch et al., ; Camargo et al., ; Romano and Zeng, ).
Due to increasing nutrient levels within ground and surface waters, a number of reviews have been published recently dealing with nitrogenous wastes in aquatic systems and their biological or ecological implications. These have included reviews of the complex active ammonia-N excretion mechanisms in crabs (Weihrauch et al., ); nitrite-N and nitrate-N uptake and elimination in aquatic animals (Jensen, ); origins and effects of elevated nitrite-N on aquatic animals emphasizing microbial involvement (Philips et al., ); nitrite-N induced physiological disruptions to aquatic animals (Jensen, ); acute nitrate-N toxicity to aquatic animals and its applications to real world scenarios and human health (Camargo et al., ; Camargo and Alonso, ); and the development of a new model to quantify increasing nitrogen loading in ecosystems from shellfish aquaculture effluent (Bouwman et al., 2011). Furthermore, Chen and colleagues have made many significant contributions over the past three decades that have greatly enhanced our understanding in this field. However, to our knowledge, a comprehensive review on the relative toxicity and physiological consequences of all three nitrogenous wastes (i.e. ammonia-N, nitrite-N and nitrate-N) to decapod crustaceans is lacking.
Decapod crustaceans, including shrimps, crabs, lobsters, and freshwater prawns and crayfish, are of high commercial value that form important fisheries globally, while the crustacean aquaculture industry is a continuously growing multi-billion dollar sector almost exclusively based on this animal group (FAO, ). In their natural habitats, decapod crustaceans are particularly susceptible to elevated nitrogenous waste exposure due to their often benthic or near benthic dwellings, generally characterized by substantially higher dissolved nitrogen levels than surface waters (Cutrofello and Durant, ; Fanjul et al., ). For example, ocean sediments that provide habitats for many benthic decapod crustaceans can reportedly have ammonia-N levels exceeding 39 mg L−1 in some areas due to high organic contents (Weihrauch et al., ). In closed aquaculture systems, the situation is more extreme than those in nature since cultured crustaceans can routinely encounter elevated levels of different nitrogenous waste forms as well as at longer durations that can result in significant economic losses.
The first section of this review will briefly discuss the implications of elevated levels of these three nitrogenous wastes in aquatic ecosystems receiving excessive effluent discharge as well as under aquaculture scenarios. Subsequently, the acute toxicity of ammonia-N, nitrite-N, and nitrate-N to decapod crustaceans and influencing factors will be discussed and compared, including a discussion on a new and more accurate protocol for conducting acute nitrate-N toxicity tests. A large part of this review will then be devoted to sub-lethal effects of nitrogenous waste exposure on the growth and various essential physiological responses of decapod crustaceans since these are the most likely scenarios and are therefore practically important. Prior to discussing the physiological consequences of decapod crustaceans to nitrogenous waste exposure, a brief background on how these physiological responses operate will be provided. Finally, the uptake, elimination and detoxification mechanisms of decapod crustaceans to elevated nitrogenous waste exposure will be discussed since these adaptive mechanisms will ultimately explain toxicity differences due to biotic and abiotic factors as well as among the different nitrogenous waste species.
Generally, ammonia-N, nitrite-N, and nitrate-N remain very low in most aquatic ecosystems (<2 mg L−1 total nitrogen) since there are natural processes that dilute these nutrients as well as being continuously removed by various aquatic plants and microbes (Philips et al., ). However, due to increasing anthropogenic activities associated with population growth and industrialization, nutrient levels are escalating within ground and surface waters globally (Curtfello and Durand, 2007; Nixon et al., ; Wear and Tanner, ). Moreover, the amount of total nitrogen discharged from rivers to coastal waters increased from 36.7 terragrams of nitrogen (TN) yr−1 in 1970 to 43.2 TN yr−1 in 2000 (Seitzinger et al., ). The origin of these nutrients are from various point and non-point sources including fertilizer run-off, livestock waste, atmospheric deposition, landfill seepage, sewage leachate, and aquaculture effluent (Seitzinger and Kroeze, ; Constable et al., ; Neal and Davies, ; Camargo et al., ; Nixon et al., ; Cutrofello and Durant, ; Sigleo and Frick, ; Wear and Tanner, ; Bouwman et al., 2011).
Due to the transient nature of nitrogenous waste, numerous studies have examined spatial and temporal variations of these concentrations within various ecosystems (Constable et al., ; Neal and Davies, ; Camargo et al., ; Nixon et al., ; Cutrofello and Durant, ; Lu et al., ; Sigleo and Frick, ; Wear and Tanner, ). For example, the input of total dissolved nitrogen discharges to some bays ranged as low as 144 tons N yr−1 to as high as 151,546 tons N yr−1depending on the proximity to human populations and seasonal rainfall (Caccia and Boyer, ). Moreover, as the most toxic nitrogenous waste species, pulses of ammonia-N levels predicted to reduce the abundance of various invertebrate species have been detected in some waterways (Cutrofello and Durant, ; Nixon et al., ).
Meanwhile, incomplete nitrification from municipal wastewater, secondary treated sewage, or landfill leachate can contain nitrite-N as high as 13 mg L−1 and such discharge can exceed those predicted to induce mortalities in aquatic animals or disrupt essential physiological processes (Phillips et al., 2002; Constable et al., ; Dave and Nilsson, ; Wear and Tanner, ). Indeed, excessive sewage discharge led to ammonia-N and nitrite-N levels reaching 2.4 and 0.7 mg L−1, respectively within some coastal waters of Australia and were implicated as the likely causes to the reduced abundance of the commercially important blue swimmer crab, Portunus pelagicus (Wear and Tanner, ). The final product of the nitrification cycle is nitrate-N, and while this is the least toxic form, it is often discharged at much higher levels (Neal and Davies, ; Camargo and Alonso, ). For example, it has been reported that nitrate-N concentrations in surface and ground waters can exceed 25 and 100 mg L−1, respectively (Camargo et al., ), while nitrate-N discharges into some rivers in England can reach as high as 2,400 kg yr−1 km−2, which is many times higher than 300 and 60 kg yr−1 km−2 for ammonia-N and nitrite-N, respectively (Neal and Davies, ).
When nitrogenous waste is discharged into estuaries/rivers, there is less of a diluting effect compared to the open ocean (Nixon et al., ; Wear and Tanner, ). This is particularly important since the life cycle of numerous crustacean species are dependent on estuary systems as vital breeding and nursery grounds (Sigleo and Frick, ). Moreover, since the composition of effluent waters often contain more than one nitrogenous waste species, potential toxic interactions with nitrate-N may be exerted on aquatic animals (Romano and Zeng, ).
Similarly, under aquaculture scenarios, the culture water in closed aquaculture systems often contain all three pollutants that routinely exceed those found in nature (Camargo et al., ; Ferreira et al., ), and to cope with this situation, high volumes of effluent water is often discharged to the surrounding environment (Shumway, ). The issues of these are twofold. Accumulated nitrogenous waste can significantly impact the productivity of cultured crustaceans and when the aquaculture effluent is discharged this can, in turn, negatively impact the residential population of the nearby ecosystems (D'Souza et al., ; Tovar et al., ; Biao et al., ; Lin et al., ; Ferreira et al., ). The origin of nitrogenous waste in intensive aquaculture systems stems from high feed inputs and subsequent metabolic by-products of the cultured animals (Nhan et al., ), which is increasingly becoming an issue as the aquaculture industry is expected to substantially increase in the coming decades, particularly in Asia (Bouwman et al., 2011; Shumway, ). For example, the mean ammonia-N, nitrite-N and total nitrogen levels can reach 2.76, 2.48, and 21.36 mg L−1, respectively, within some shrimp ponds (Burford et al., ). Meanwhile, persistent levels of nitrite-N as high as 4.61 mg L−1 have been reported (Chen et al., ), while in closed recirculating systems, nitrate-N levels of 150 mg L−1 can be reached (Kaiser and Schmitz, ). Even in aquaculture systems with fully established biofilters or other natural removal processes (e.g., algal growth), dynamic changes in nitrogenous waste levels sufficient to induce physiological changes/adaptations can still occur daily (Nhan et al., ). This likely explains the bulk of research in this area within the context of aquaculture (e.g., Chen and Cheng, ; Jensen et al., ; Hong et al., ).
Acute toxicity tests represents a standard method for quantifying and comparing relative toxicities of pollutants and is determined by measuring the concentration that kills 50% of the tested organisms, called the LC50 value (APHA, [ 3]). Using probit analysis, LC50 values are generally presented after 12 to 24 hours of exposure to the pollutant, and then at 12-hour intervals up to 96 hours. In addition to providing comparable data to other species or pollutants, these values are particularly useful when examining relationships between the toxicant and other abiotic or biotic variables. It is generally known that ammonia-N often has a lower LC50 value than either nitrite-N or nitrate-N, and is therefore more toxic (e.g., Chen and Lei, ; Chen et al., ,; Meade and Watts, ).
Graph: Figure 1 The NH4+ to NH3 ratios at different pH, temperature, and salinities with higher ratios favoring the prevalence of the NH4+ form. Values calculated based on Whitfield ().
In aquatic environments, ammonia-N exists in two forms as NH4+ and NH3, with the latter form being more toxic since it can more easily diffuse across lipid bilayers (i.e., the gills of crustaceans) and into the hemolymph (Weihrauch et al., ). The NH4+ to NH3 ratios are dependent on salinity, temperature, and, to a greater extent, pH (Figure 1) and each of these variables influences ammonia-N toxicity in different ways (Lin and Chen, 2001). Although few studies have investigated the influence of pH on ammonia-N toxicity to crustaceans, the results are consistent showing a higher pH increases ammonia-N toxicity. This is likely due to the NH3 form becoming more prevalent (Armstrong et al., [ 5]; Noor-Hamid et al., ). However, despite salinity and temperature having a much smaller effect on NH4+ to NH3 ratios, they can still greatly affect ammonia-N toxicity (Table 1). For example, an inverse relationship exists between salinity and ammonia-N toxicity, as observed for red tail shrimp, Penaeus penicillatus (Chen and Lin, ), fleshy shrimp, Penaeus chinensis (Chen and C. Y. Lin, 1992a), white Pacific shrimp, Litopenaeus vannamei (Lin and Chen, 2001; Li et al., 2007), green tiger shrimp, Penaeus semisulcatus (Kir and Kumlu, ), white shrimp, Litopenaeus schmitti (Barbieri, [ 8]), and blue swimmer crab, Portunus pelagicus (Romano and Zeng, ). In the case of the latter, such a relationship was linked with detecting higher hemolymph ammonia-N levels in P. pelagicus at lower salinities compared to those at higher salinities likely due to the utilization of different coping mechanisms (Romano and Zeng, ), which will be discussed in more detail later in the article.
Table 1 The 96-hour LC50 values (mg L−1) of ammonia-N (NH4+ and NH3) and unionised ammonia (NH3-N), unless otherwise stated, from various crustacean species
|Species||Life stage||Weight (g)||Salinity||Temperature||Ammonia-N (mg L−1)||NH3-N (mg L−1)||Reference|
| P. monodon||Postlarvae||34||29.5°C||11.51||1.04||Chin and Chen (1987)|
| P. monodon||Postlarvae||2.2||34||25°C||37.40||1.69||Allan et al. (1990)|
| M. macleayi||Postlarvae||2.0||34.5||25°C||26.30||1.39||Allan et al. (1990)|
| M. ensis||Juvenile||0.01||25||25°C||35.59||0.87||Nan and Chen (1991)|
| P. chinensis||Juvenile||0.61||10||25°C||28.18||1.94||Chen and C. Y. Lin (1992a)|
| P. paulensis||Juvenile||5.45||28||25°C||38.72||1.10||Ostrensky and Wasielesky (1995)|
| P. setiferus||Postlarvae||25||28°C||9.38 (48 hours)||1.21 (48 hours)||Alcaraz et al. (1999)|
| P. semisulcatus||Juvenile||1.60||39||14°C||55.84||1.92||Kir et al. (2004)|
| P. semisulcatus||Postlarvae||0.27||15||25°C||7.07||0.34||Kir and Kumlu (2006)|
| P. penicillatus||Juvenile||0.40–0.69||34||20°C||38.72||1.11||Chen and Lin (1991)|
| L. vannamei||Juvenile||15||23°C||24.39||1.20||Lin and Chen (2001)|
| L. schmitti||Juvenile||1.50||5||20°C||19.12||0.69||Barbieri (2010)|
| H. americanus||Postlarvae||30||20°C||144||2.36||Young-Lai et al. (1991)|
| E. sinensis||Zoea I||NR||25||22°C||5.70 (72 hours)||0.20 (72 hours)||Zhao et al. (1997)|
|Zoea II||NR||25||22°C||7.60 (72 hours)||0.27 (72 hours)|
|Juvenile||0.06||FW||22°C||45.30 (72 hours)||1.29 (72 hours)|
| E. sinensis||Juvenile||3.2||FW||20°C||119.6||1.31||Hong et al. (2007)|
| N. granulata||Adult||8.27||5||20°C||141.50||0.55–1.13A||Rebelo et al. (1999)|
| P. pelagicus||Juvenile||0.002||30||28°C||23.10||1.65||Romano and Zeng (2007c)|
| S. serrata||Zoea I||32||28.3°C||50.00 (48 hours)||3.27 (48 hours)||Neil et al. (2005)|
|Zoea II||32||28.3°C||33.56 (48 hours)||2.18 (48 hours)|
|Zoea III||32||28.3°C||37.52 (48 hours)||2.40 (48 hours)|
|Zoea IV||32||28.3°C||43.54 (48 hours)||2.83 (48 hours)|
|Zoea V||28||28.3°C||47.13 (48 hours)||3.09 (48 hours)|
|Megalopae||28||28.3°C||20.63 (48 hours)||1.35 (48 hours)|
| S. serrata||Juvenile||0.373||30||28°C||95.35||6.81||Romano and Zeng (2007d)|
|AThe NH3-N values were not calculated, and the values presented in this table are calculated based on a pH of 7.00 used in the study and on a speculative temperature range of 20–30°C.|
|NR: Not reported. FW: Freshwater.|
Ammonia-N toxicity can also depend on the developmental stage of crustaceans, and while there are sometimes large tolerance differences between larval stages, generally post-larvae or juveniles have higher ammonia-N tolerances that are positively correlated with size (Ostrensky and Wasielesky, ; Zhao et al., ; Romano and Zeng, ). For example, the 72-hour LC50 value of zoea I larvae of the Chinese mitten crab, Eriocheir sinensis, was 5.7 mg L−1 and increased substantially to 45.3 mg L−1 at the juvenile stage (Zhao et al., ). Similarly, there was a greater than twofold increase in the ammonia-N tolerance from newly settled P. pelagicus first stage juveniles crabs compared to when the 7th juvenile crab stage was reached (Romano and Zeng, ). In the latter case, more severe gill damage was found in younger crabs than older crabs exposed to the same ammonia-N level, which will be discussed in more detail in the "Histological Damage" section. Neil et al. () found no relationship between ammonia-N toxicity and the larval stages of the mud crab, Scylla serrata, although the juvenile crab stage had more than a twofold higher tolerance (Romano and Zeng, ). An exception has been found in Metapenaeus ensis juveniles as their ammonia-N tolerance was often lower than the less developed larval stages (Chen and Nan, ; Nan and Chen, ; see Table 1).
Despite nitrite-N often reaching high levels in sewage treatment plants or in aquaculture systems due to inadequate nitrification, acute nitrite-N toxicity to crustaceans is substantially less studied compared with those of ammonia-N. Nevertheless, available information appears to indicate that pH, salinity, and developmental stages are major influential factors.
Similar to ammonia-N, nitrite-N exists in two forms: NO2− ions and nitrous acid (HNO2). While HNO2− is often measured in the μg L−1 scale, compared to mg L−1 of NO2−, this is still significant since HNO2− can easily diffuse across the gill membranes of aquatic animals, thus making this form more toxic. While the ratio of these nitrite forms are greatly influenced by pH, to the best of our present knowledge, no acute nitrite-N toxicity tests at different pH levels are available on decapod crustaceans. However, correlations between nitrite-N toxicity and pH have been demonstrated with fish (Russo et al., ; Huey et al., ) and a similar trend likely exists for crustaceans since Chen and Cheng (2000) found that both nitrite-N influx and hemolymph nitrite-N levels increased in black tiger prawn, Penaeus monodon, at decreasing pH.
Despite relatively few studies on the effects of salinity or Cl− levels, results show consistent trends (Table 2). For instance, in freshwater (<0.5 mg L−1 Cl−), a relatively low 96-hour LC50 value of 6.1 mg L−1 was measured for crayfish Procambarus simulans, and when the Cl− concentration was increased to 300 mg L−1, no mortalities were observed and hence LC50 values could not be computed (Beitinger and Huey, [ 9]). Similarly, higher salinities (and therefore higher Cl− levels) significantly reduced nitrite-N toxicity to both P. penicillatus (Chen and Lin, ) and L. vannamei (Lin and Chen, ). In contrast to the above findings, salinity was shown to have no effect on the nitrite-N LC50 values to S. serrata larvae (see Table 2) (Seneriches-Abiera et al., ). However, if these larvae possess underdeveloped osmoregulatory abilities than their older counterparts, as observed in other crabs (Charmantier et al., ; Anger and Charmantier, [ 4]), reduced nitrite-N/Cl− uptake may explain this finding (see the "Uptake, Elimination, and Detoxification Mechanisms: Nitrite" section for further discussion) although this requires further investigation.
Table 2 The various 96 hours LC50 values (mg L−1) of NaNO2-N from different decapod crustacean species
|Species||Size or life stage||Culture condition||pH||LC50 (mg L−1)||Reference|
|P. simulans||Unknown||FW (< 0.5 mg L−1 Cl−)||Unknown||6.1 (96 hours)||Beitinger and Huey (1981)|
|FW (300 mg L−1)||n/aA|
|P. clarkii||Juveniles (0.5–12.4 g)||FW (22 mg L−1 Cl−)||7.9–8.3||28.0 mg L−1 (96 hours)||Gutzmer and Tomasso (1985)|
|Juveniles (8.4 g)||FW (22 mg L−1 Cl−)||7.9–8.3||≈30 mg L−1 (96 hours)|
|Juveniles (7.1 g)||FW (100 mg L−1 Cl−)||7.9–8.3||≈80 mg L−1 (96 hours)|
|M. rosenbergii||10–14 day old larvae||12||7.98–8.22||1.8 (96 hours)||Armstrong et al. (1976)|
|M. rosenbergii||2.02–2.93 g||FW (15 mg L−1 Cl−)||7.83||8.49 (96 hours)||Chen and Lee (1997)|
|FW (24 mg L−1 Cl−)||7.83||11.21 (96 hours)|
|FW (34 mg L−1 Cl−)||7.83||12.87 (96 hours)|
|M. nipponese||1.05 g||FW||13.3 (96 hours)||Wang et al. (2004)|
|M. malcolmosonii||Juveniles (10–15 g)||FW||7.5–8.0||3.14 (96 hours)||Chand and Sahoo (2006)|
|P. leniusculus||Adult (33.04 g)||AFWMB (17.7 mg L−1 Cl−)||7.6||31 (48 hours)||Harris and Coley (1991)|
|A. leptodactylus||27.15 g||FW (35 mg L−1 Cl−)||7.21||29.43 (48 hours)||Yildiz and Benli (2004)|
|FW (100 mg L−1 Cl−)||49.20 (48 hours)|
|P. monodon||Nauplius||34||7.98–8.22||5.00 (24 hours)||Chen and Chin (1988)|
|Zoea||34||7.98–8.22||13.20 (24 hours)|
|Mysis||34||7.98–8.22||20.65 (24 hours)|
|Post-larvae||34||7.98–8.22||61.87 (24 hours)|
|P. monodon||Adolescents (4.87 g)||20||7.57||171 (96 hours)||Chen et al. (1990b)|
|P. chinensis||Juveniles (0.36 g)||33||7.94||37.71 (96 hours)||Chen et al. (1990a)|
|P. penicillatus||Juveniles (0.40–0.69 g)||25||8.10||38.52 (96 hours)||Chen and Lin (1991)|
|35||8.10||40.86 (96 hours)|
|M. ensis||3rd nauplii||33||8.20||31.29 (24 hours)||Chen and Nan (1991)|
|2nd zoea||33||8.20||16.05 (24 hours)|
|2nd mysis||33||8.20||47.60 (24 hours)|
|1st postlarvae||33||8.20||70.06 (24 hours)|
|P. setiferus||Postlarvae||25||8.42||248.84 (48 hours)||Alcaraz et al. (1999)|
|L. vannamei||Juveniles (3.96 g)||15||8.02||76.5 (96 hours)||Lin and Chen (2003)|
|25||8.02||178.3 (96 hours)|
|35||8.02||321.7 (96 hours)|
|S. serrata||Zoea I||25–35C||8.32–8.64||41.58 (96 hours)||Seneriches-Abiera et al. (2007)|
|Zoea II||25–35||8.32–8.64||63.04 (96 hours)|
|Zoea III||25–35||8.32–8.64||25.54 (96 hours)|
|Zoea IV||25–35||8.32–8.64||29.98 (96 hours)|
|Zoea V||25–35||8.32–8.64||69.93 (96 hours)|
|ANo mortalities to calculate LC50 values.|
|BArtificial freshwater media had 2.5 mM CaSO4, 0.25 mM KHCO3, and 0.5 mM MgSO4|
|CSalinities had no affect on LC50 values and these values were averaged in the study.|
Only two experiments have been performed on nitrite-N toxicity at different developmental stages and with contrasting results (Chen and Chin, ; Seneriches-Abiera et al., ). Chen and Chin () found with the progression of larval development from nauplii to postlarvae, the nitrite-N tolerance of P. monodon increased over 12-fold (a 96-hour LC50 value of 5.00 mg L−1 to 61.87 mg L−1). In contrast, each zoeal stage of S. serrata larvae had markedly different nitrite-N tolerances and no clear link to their developmental stages was shown (Seneriches-Abiera et al., ; see Table 2).
Nitrate (NO3−) is the final step in the nitrification process, and since this has a much lower toxicity compared with ammonia-N and nitrite-N, it is not surprising that this is the least studied among the three nitrogenous species. To the best of our knowledge, information on acute nitrate-N toxicity is limited to only a few decapod species that include the redclaw crayfish, Cherax quadricarinatus (Meade and Watts, ), P. monodon (Tsai and Chen, ; Romano and Zeng, ), P. pelagicus and S. serrata (Romano and Zeng, ; Table 3).
Table 3 The various 96-hour LC50 values (mg L−1) of NaNO3-N from different decapod crustaceans
|Cherax quadricarinatus||Freshwater||1,000||Meade and Watts (1995)|
|Penaeus monodon||15 without K+||1,411||Romano and Zeng (2009)|
|15 with K+||1,878|
|25 without K+||2,020|
|25 with K+||2,297|
|35 without K+||2,213|
|35 with K+||2,337|
|Portunus pelagicus||30 without K+||3,355||Romano and Zeng (2007)|
|30 with K+||4,132|
|Scylla serrata||30 without K+||3,601||Romano and Zeng (2007)|
|30 with K+||4,339|
Table 4 Required volumes of pre-adjusted saline water and required volume additions of the NaNO3-N stock solution to create the desired NaNO3-N concentration for each 10 L final test solution at salinities of 15 and 25. The 35 salinity treatment is not included since, even at the highest NaNO3-N concentrations of 5000 mg L−1, the salinity change would equal only 1.0. Data obtained from Romano and Zeng (2009b).
|Volume of pre-adjusted||Volume of NaNO3-N|
|Desired NaNO3-N||saline water required to||stock solution|
|concentration||dilute the stock solution||(salinity 36) required|
|15 salinity treatment|
| 1000 mg L−1||9.0 L at salinity 13||1.0 L|
| 1500 mg L−1||8.5 L at salinity 11.5||1.5 L|
| 2000 mg L−1||8.0 L at salinity 10||2.0 L|
| 2500 mg L−1||7.5 L at salinity 8||2.5 L|
| 3000 mg L−1||7.0 L at salinity 6||3.0 L|
| 3500 mg L−1||6.5 L at salinity 4||3.5 L|
|25 salinity treatment|
| 1000 mg L−1||9.0 L at salinity of 24||1.0 L|
| 1500 mg L−1||8.5 L at salinity of 23||1.5 L|
| 2000 mg L−1||8.0 L at salinity of 22||2.0 L|
| 2500 mg L−1||7.5 L at salinity of 21||2.5 L|
| 3000 mg L−1||7.0 L at salinity of 20||3.0 L|
| 3500 mg L−1||6.5 L at salinity of 19||3.5 L|
Nevertheless, it should be highlighted that there is evidence suggesting that nitrate-N may be even less toxic that previously reported, which is likely applicable to all estuarine/marine animals. This is due to inherent problems with detecting LC50 values of nitrate-N that alter the ionic composition of the test solutions, leading to osmoregulatory stress unrelated to nitrate-N and, consequently, mortality bias (Romano and Zeng, ). To understand this, the process of conducting acute toxicity tests will be briefly explained.
The first step of an acute toxicity test is creating a stock solution, normally at a concentration of 10,000 mg L−1 that is subsequently diluted to create a range of the desired test concentrations. In the case of nitrate-N, when NaNO3 is added to distilled water to create the stock solution it will only contain NO3− and Na+ ions. Due to the lower toxicity of nitrate-N, high volumes of stock solution are normally required. For example, LC50 values for aquatic animals are generally >1,000 mg L−1, and to make a 10-L test solution of 1,000 mg L−1 nitrate-N, 1 L of the 10,000 mg L−1 is added to 9 L of saltwater (Table 4). These large stock solution dilutions, with only Na+ and NO3− ions, will inevitably have a significant impact on the ionic composition of the final test solution, such as increasing Na+/K+ ratios. Such disrupted ion ratios, which are compounded at lower salinities (Figure 2), is well known to induce osmoregulatory stress and subsequent death in aquatic animals regardless of their osmoregulatory abilities (Romano and Zeng, ; Romano and Zeng, ).
A simple method to mitigate these ratio changes is through K+ additions to the nitrate-N stock solution, at the same level as those naturally found in the test solution salinity, which significantly reduced mortality bias in P. pelagicus, S. serrata (Romano and Zeng, ) and P. monodon (Romano and Zeng, ). While this can reduce osmoregulatory stress, it is cautioned that nitrate-N toxicity comparisons of the same species but at different salinities may not be entirely accurate (Romano and Zeng, ). Nevertheless, since nitrogenous waste toxicity is often linked with hemolymph accumulation, experiments on hemolymph nitrate-N levels of crustaceans, when subjected to different nitrite-N and salinity combinations, may provide insight into toxicity trends.
Often nitrogenous waste levels in either nature or aquaculture systems rarely exceed the respective LC50 values of crustaceans. However, experiments on sub-chronic/chronic nitrogenous waste exposure on the growth and/or feeding rates of crustaceans are relatively limited. One of the reasons may be due to both the relative ease and more standardized method of conducting shorter acute toxicity test to measure LC50 values, where environmental conditions (e.g., temperature, pH, salinity, and nitrogenous waste levels) can be tightly controlled. Moreover, LC50 values are often used to calculate "safe" concentrations by multiplying 96-hour LC50 values by an empirical factor of 0.1 (i.e., 10% of 96-hour LC50 values) that is then used to predict the concentration that is not only non-lethal but allows the animal to thrive (Sprague, ).
Graph: Figure 2 The altered Na+/K+ ratios (mg L−1) of the control (natural seawater) and at increasing NaNO3-N concentrations when K+ was added, at the same K+ level found in natural seawater at 30, and when K+ was not added at the salinity levels of (A) 15, (B) 25, and (C) 35. The broken line with triangles and squares represent Na+/K+ ratios without and with added K+, respectively, and the solid line with diamonds represent Na+/K+ ratio of 27:1.
However, there is evidence indicating that such "safe" concentrations of ammonia-N and nitrite-N are a substantial overestimation, which will be briefly discussed later. Furthermore, various laboratory experiments analyzing sensitive stressor endpoints, such as gas exchange, acid/base balance, immunological responses, osmoregulation, and histopathology, often reveal significant disruptions/alterations at relatively low concentrations. Therefore investigations on sublethal exposure or incorporating physiological measurements during acute toxicity tests can provide important and more comprehensive information, particularly for environmental risk assessments and aquaculture management (Dahl et al., ).
In an experiment with Metapenaeus japonicus, Chen and Kou () revealed that the maximum acceptable toxicant concentration (MATC) of ammonia-N (measured as percentage weight gains) within 30 and 60 days was 5 mg L−1 and <5 mg L−1, respectively while Chen and C. Y. Lin () reported a MATC of 6 mg L−1 ammonia-N for P. penicillatus. For giant freshwater prawn, Macrobrachium rosenbergii, exposure to ammonia-N levels as low as 0.5 mg L−1 significantly reduced their growth rates (measured as weight gain/day) as well as reduced feed intake after 60 days (Naqvi et al., ), while the lowest MATC were obtained for school prawn, Metapenaeus macleayi and P. monodon, of 0.3 and 0.21 mg L−1 ammonia-N, respectively, after 30 days (Allan et al., [ 2]).
These experiments show substantial discrepancies between theoretical and real "safe" concentrations. For example, the 96-hour LC50 value of P. penicillatus at a salinity of 34 was 38.72 mg L−1 ammonia-N, thus providing a "safe" level of 3.87 mg L−1 ammonia-N (Chen and Lin, ); however, 3 mg L−1 ammonia-N was sufficient to reduce the growth of this same species within 14 days (Chen and C. Y. Lin, ). Moreover, based on the 96-hour LC50 values from Allan et al. ([ 2]), the "safe" concentrations for M. macleayi and P. monodon would theoretically be 2.6 and 3.7 mg L−1 ammonia-N, although in the same experiment, the MATC were 0.3 and 0.21 mg L−1 ammonia-N for M. macleayi and P. monodon, respectively (Allan et al., [ 2]). In the case of M. rosenbergii, an ammonia-N concentration of 0.5 mg L−1 ammonia-N was sufficient to reduce their growth (Naqvi et al., ), implying a 96-hour LC50 value of ≥5 mg L−1 ammonia-N, which may be unrealistically low. These discrepancies, which can be over tenfold from the "safe" concentration, underscores the importance of conducting chronic/sublethal experiments, particularly when in conjunction with investigating the various causes implicated as for reducing their growth/physiological conditions, which will be extensively discussed later in the article.
Similarly to ammonia-N, nitrite-N can significantly affect the growth of crustaceans, albeit typically higher concentrations are necessary (Koo et al., ). For example, while ammonia-N levels of 50 mg L−1 significantly reduced the growth of the tiger crab, Orithiya sinica, when compared to the control, a nitrite-N concentration of 150 mg L−1 was necessary to induce a similar significant growth reduction after 30 days (Koo et al., ). Interestingly, both ammonia-N and nitrite-N significantly accelerated the molting of O. sinica (Koo et al., ) and, similarly, 4 mg L−1 nitrite-N increased the molting frequency of P. monodon juveniles, although the size increase at each molt was reduced causing lower growth after 20 days (Chen and Chen, ). In the case of the later, the "safe" levels for P. monodon postlarvae and juveniles would theoretically be 6.1 and 17.1 mg L−1 nitrite-N, respectively (Chen and Chin, ; Chen et al., ), which is a substantial overestimation (Table 2).
Nitrite-N also affected the molting of P. pelagicus juveniles, but rather than increasing molting frequency, nitrite-N levels from 3 to 16 mg L−1 significantly increased "molt death syndrome" (Romano and Zeng, ). This was characterized by a mortality resulting from the inability of the crab to completely remove their old exoskeleton during molting and diagnosed as an incomplete disengagement of the previous shell from a dead crab. Furthermore, P. pelagicus had a high sensitivity to nitrite-N since the lowest tested level of 3 mg L−1 led to significantly lower growth after 20 days of exposure (Romano and Zeng, ). This is particularly noteworthy considering their high ammonia-N tolerance (Romano and Zeng, ) indicating that a high tolerance to one nitrogenous form may not necessarily correspond to another.
For M. rosenbergii larvae, it has been shown that the earlier larval stages are more susceptible than more developed larvae (Mallasen and Valenti, ). Mallasen and Valenti () measured the cumulative survival and growth of M. rosenbergii from larvae I to VIII (phase 1) and larvae VIII to postlarve (phase 2). It was shown that nitrite-N levels above 2 mg L−1 significantly reduced the survival and growth of M. rosenbergii in phase 1, while for the survival and growth to be significantly reduced nitrite-N levels of 16 and 8 mg L−1, respectively, were required (Mallasen and Valenti, ).
The effects of nitrate-N on the growth and development to crustaceans are limited, and to our knowledge, restricted to M. rosenbergii over progressive larval stages (Mallesen et al., 2004) and L. vannamei juveniles (Kuhn et al., ). In a series of experiments using successively higher nitrate-N concentrations, Mallesen et al. (2004) indicated that nitrate-N levels of 180 mg L−1 would not be a significant stressor to M. rosenbergii larvae. For L. vannamei, two separate experiments were conducted at different nitrate-N concentrations of 220, 435, and 910 mg L−1 and then at different salinities of 2, 9, and 18 at the same nitrate-N concentration of 440 mg L−1 (Khun et al., 2010). In the first experiment, it was shown that 440 and 910 mg L−1 nitrate-N significantly reduced the growth and survival of L. vannamei, respectively from two weeks onwards while lower salinities at 440 mg L−1 had a greater impact on prawn survival and growth than at higher salinities. Much was made regarding osmoregulatory stress increasing the energetic demands of L. vannamei as an explanation to this finding (Khun et al., 2010), however it remains unclear whether the altered ionic profile to the culture water was also a contributing factor.
The mechanisms of gas exchange and acid/base balance are inter-related with the enzyme, carbonic anhydrase (CA), being responsible for both these processes. In crustaceans, CO2 is transported within the hemolymph via HCO3−, where basolaterally located CA catalyses this back to CO2 (+ H2O) to be passively excreted to the environment. For cytoplasmically located CA, this hydrates CO2 back to H+ and HCO3−, which can be used as an anti-porter for Na+ and Cl− via apically located Na+/H+ (or V-type H+-ATPase) and Cl−/HCO3− transporters, respectively (see Freire et al. [] for review). In crustaceans, oxygen is transported through the hemolymph via a copper (Cu) containing respiratory protein hemocyanin and, once this binds to one O2 molecule, it becomes oxyhemocyanin (Jensen, ; Cheng and Chen, ). Hemocyanin levels, oxyhemocyanin/deoxyhemocyanin ratios, and/or oxygen consumption are commonly measured as indicators of gas exchange in crustaceans, although to our knowledge, none have measured the role or effects of CA activity during nitrogenous waste exposure.
Direct evidence of ammonia-N induced disruptions to gas exchange has been somewhat paradoxical. Generally, exposure to elevated ammonia-N levels leads to decreased oxyhemocyanin in numerous crustaceans, including P. monodon (Chen and Cheng, ), M. japonicus (Chen et al., 1994) and L. vannamei (Racotta and Hernández-Herrera, ) while, on the other hand, ammonia-N has been shown to increase oxygen consumption for M. japonicus (Chen and Lai, ), P. chinensis (Chen and J. N. Lin, ; Chen and C. Y. ), L. vannamei (Racotta and Hernández-Herrera, ) and Macrobrachium nipponense (Wang et al., ). Racotta and Hernández-Herrera () noted such a phenomenon, especially since compensatory anaerobic metabolism did not appear to elevate, evidenced by no significant increase in hemolymph lactate levels for L. vannamei. It was nevertheless suggested that oxygen transport in crustaceans would likely be inhibited, and an explanation for the unaffected lactate levels was the result of the stressful nature in measuring oxygen consumption, thus, masking these effects (Racotta and Hernández-Herrera, ). This seems reasonable since elevated lactate levels have been reported for ammonia-N exposed E. sinensis (Hong et al., ), while reduced dissolved oxygen concentrations increases ammonia-N toxicity for P. monodon juveniles (Allan et al., [ 2]) and the lobster Homarus americanus adults (Young-Lai et al., ). However, causes for increased oxygen consumption for M. japonicus and P. chinensis require more investigation, and measuring hemolymph CO2 production, lactate levels and/or CA activity may provide additional insight. It is perhaps worthy to note that within one hour of exposure to 5.4 mg L−1 ammonia-N, rainbow trout, Oncorhynchus mykiss, had significantly lower gill CA activity, although the exact mechanisms were not fully elucidated (ArasHisar et al., [ 7]).
It is well known that elevated nitrite-N severely disrupts the oxygen carrying capacity in fish when this nutrient enters the blood cells and oxidizes hemoglobin (FeII) to methemoglobin (FeIII), thus becoming incapable of binding to oxygen (Williams et al., ). However, there are conflicting reports on whether a similar process occurs for crustaceans. Cheng and Chen () suggested that, for M. japonicus, hemolymph nitrite-N would oxidize hemocyanin to methemocyanin or that deoxyhemocyanin would be unable to be oxygenated. This suggestion was supported by detecting significantly lower oxyhemocyanin and proteins levels within the hemolymph of M. rosenbergii (Chen and Lee, ) and M. japonicus juveniles (Cheng and Chen, ) as well as reductions to hemolymph oxyhemocyanin and protein along with increased hemolymph deoxyhemocyanin in P. monodon (Cheng and Chen, ). Furthermore, there were also hemolymph pO2 increases to M. rosenbergii (Chen and Lee, ), as well as increased hemolymph pO2 and reduced oxygen affinities (expressed as P50) for both P. monodon (Cheng and Chen, ) and M. japonicus (Cheng and Chen, ).
However, there is evidence suggesting gas exchange is relatively unaffected in other crustaceans (Jensen, ; Mallasen and Valenti, ). Despite measuring significantly lower hemocyanin in nitrite-N exposed freshwater crayfish, Astacus astacus, Jensen () stated that their oxygen carrying capacity was relatively unaffected evidenced by stable hemolymph lactate levels throughout elevated nitrite-N exposure. Moreover, Tahon et al. () demonstrated that methemocyanin would only occur at especially high nitrite-N levels and low pH, and such extreme conditions would likely be the cause of mortality rather than potential nitrite-N induced gas exchange disruptions. In addition, M. rosenbergii staged II larvae exposed to elevated nitrite-N had significantly higher respiration rates, as measured by oxygen consumption, indicating the gas exchange mechanisms accelerated (Mallasen and Valenti, ).
Conflicting reports have also been made regarding hemolymph pH. It has been shown that exposure to elevated nitrite-N concentrations significantly reduced hemolymph pH in the signal crayfish, Pacifastacus leniusculus (Harris and Coley, ), M. rosenbergii (Chen and Lee, ), P. monodon (Cheng and Chen, ) and M. japonicus (S. Y. Cheng and Chen, ). These findings of hemolymph acidosis were suggested to be linked with increased CO2 levels (S. Y. Cheng and Chen, ). On the other hand, for A. astacus exposed to elevated nitrite-N, the hemolymph pH increased and was suggested to be the result of hyper-ventilation leading to lower hemolymph CO2 levels (Jensen, ). Clearly more research should continue in this area, such as measuring hemolymph lactate levels and CA activity, which may help elucidate some of these findings.
Crustaceans only possess an innate immunity but no acquired immunity, and therefore one of the most important immunological responses for crustaceans is the production/release of hemocytes into the hemolymph (Johansson et al., ; Ellis et al., ). This is often measured as total hemocyte counts (THC); these circulating cells are responsible for the recognition of foreign particles and subsequent phagocytosis and cytotoxicity that is stimulated by the production of the enzyme prophenoloxidase (proPO). Cytotoxicity can be measured by their respiratory burst as nitroblue tetrazolium (NBT), which releases reactive oxygen species (ROS) such as superoxide anion (O2−), hydrogen peroxide (H2O2) and free radicals to kill pathogens (Johansson et al., ). Superoxide dismutase (SOD) is also measured to indicate cellular oxidant/antioxidant balance since this pathway plays an important role when protecting against cellular free radical damage (Hong et al., ).
Studies have shown that the immunological responses, particularly hemolymph THC, of crustaceans following elevated ammonia-N exposure are highly species-specific. For example, following elevated ammonia-N exposure, hemolymph THC significantly decreased in M. japonicus (Jiang et al., ), Chinese mitten crab, Eriocheir sinensis (Hong et al., ), Indian spiny lobster, Panulirus homarus (Verghese et al., ), and L. schmitti (Rodríguez-Ramos et al., ); there was no significant change for M. rosenbergii (Cheng et al., ) and white shrimp Litopenaeus vannamei (Liu and Chen, ), and a significant increase for P. pelagicus (Romano and Zeng, ).
While the responses of THC in crustaceans to ammonia-N are inconsistent among various crustacean species, there is evidence indicating THC and associated phagocytic activity correspond to pathogen susceptibility. Liu and Chen () performed a comprehensive study on the effects of ammonia-N to the immunology and resistance of L. vannamei to the bacteria Vibrio alginolyticus. It was shown that despite an ammonia-N exposure of 21.60 mg L−1 for 168 hours, there was no significant change to THC, hyaline cells, granular cells, or phagocytic activity when compared to the control, although other immunological responses were affected. After 168 hours of exposure to 5.24 and 11.21 mg L−1, PO activity and clearance activity significantly decreased, respectively, while at 21.60 mg L−1 at 168-hour significantly increased and decreased respiratory burst/O2− production and SOD, respectively (Liu and Chen, ). Interestingly, despite these findings, ammonia-N exposure had no effect on the mortality rates of L. vannamei to V. alginolyticus challenge (Liu and Chen, ). Furthermore, in two separate experiments on M. rosenbergii, it was revealed that elevated ammonia-N exposure increased the susceptibility to both Enterococcus (W. Cheng and Chen, ) and Lactococcus garvieae (Cheng et al., ). While reduced immunological responses, measured as phagocytic activity, clearance efficiency (Cheng et al., ), and PO activity (W. Cheng and Chen, ), were noted as the likely causes to increase pathogenic susceptibility, Cheng and Chen () suggested increased respiratory burst, thus increasing O2− production, were also contributing factors.
Finally, in a time course experiment, Jiang et al. () measured a variety of immunological indicators that included proPO, THC, hemocyte phagocytosis, plasma proteins alkaline phosphatase (ALP), and nitric oxide synthase (NOS) at regular intervals from 0 to 198 hours. M. japonicus were exposed to 5 mg L−1 ammonia-N with and without challenge to white spot syndrome virus (WSSV) and those only challenged with WSSV. While many of these immunological responses fluctuated widely throughout 198 hours of exposure, there was a general trend of an immunological compromise for prawns following ammonia-N exposure that was more severe for those challenged with both WSSV and ammonia and WSSV (no ammonia-N challenge) than only ammonia-N (Jiang et al., ).
A number of experiments have been performed measuring the effects of nitrite-N to the immunological responses, susceptibility to pathogens, or both to crustaceans (Cheng et al., ; Tseng and Chen, ; Yildiz and Benli, ; Wang et al., ; Chand and Sahoo, ; Romano and Zeng, ; Xian et al., ). Generally, it has been shown that elevated nitrite-N levels depress immunological function and/or increase oxidative damage, thus increasing pathogenic susceptibility of crustaceans. It has been demonstrated that exposure to 1.59 mg L−1 nitrite-N for seven days caused a significant reduction to phagocytic activity and clearance efficiency, while respiratory burst significantly increased in M. rosenbergii (Cheng et al., ). Although there was no significant effect to THC or PO, elevated nitrite-N levels caused a significant increase to the susceptibility of M. rosenberggii to L. garvieae, suggested to be the result of compromised immunity and increased production of O2− (Cheng et al., ). Wang et al. () also speculated that imbalances between antioxidant protection and free radical production to M. rosenbergii when subjected to elevated nitrite-N levels likely contributes to nitrite toxicity. With a closely related freshwater prawn, M. malcolmsonii, nitrite-N levels of 0.063 and 0.314 mg L−1 significantly reduced PO activity and THC, respectively, and when challenged with Aeromonas hydrophila, those prawns subjected to elevated nitrite-N (0.314 mg L−1) experienced significantly higher mortality rates (Chand and Sahoo, ). Similarly, L. vannamei challenged with V. alginolyticus had higher mortality rates when exposed to 4.94 to 19.99 mg L−1 nitrite-N and it was suggested that a combination of significantly lower PO activity and THC and significantly higher respiratory burst, thus increasing O2− production, contributed to this finding (Tseng and Chen, ).
A more detailed oxidative profile and effects to hemocytes were investigated on P. monodon when Xian et al. () measured the THC, ROS, DNA damaged cells, and apoptotic cell ratios at increasing nitrite-N levels over time. It was demonstrated that exposure to 10 and 20 mg L−1 nitrite-N for 48 and 12 hours, respectively, caused a significant increase to the production of ROS, which included O2−, H2O2, hydroxyl (HO−) and peroxyl (ROO−), and was correlated with increased DNA damaged hemocytes and apoptotic ratios (Xian et al., ). Xian et al. () provided evidence for reasons causing nitrite-N induced THC reductions in P. monodon and may similarly apply to other crustacean species.
Immunological compromise may be mitigated by higher Cl− levels in crustaceans. For the crayfish, Astacus leptodactylus, exposed from 9 to 25 mg L−1 nitrite-N for 48 hours led to significantly lower THC, compared to the control. However, when NaCl was present (100 mg L−1), the same concentrations of nitrite-N significantly increased the THC of A. leptodactylus after 48 hours (Yildiz and Benli, ). Yildiz and Benli () suggested that since Cl− decreased hemolymph nitrite-N uptake from the environment, increased THC indicated A. leptodactylus was under less stress compared those under Cl− free water conditions. Similarly, in a longer term experiment, Romano and Zeng () showed that P. pelagicus significantly increased THC after 20 days of exposure to 3 to 24 mg L−1 nitrite-N at a salinity of 30. However, considering this was only measured at one salinity, research should continue to determine whether immunological responses and pathogenic susceptibility change at different salinities, which may further explain contributing factors influencing nitrite-N toxicity in crustaceans.
The majority of decapod crustaceans osmoregulate to some degree and this is an essential mechanism for survival, growth, and maintenance of many physiological functions (see Romano and Zeng, , for review). For crustaceans, osmoregulatory strength is often measured by evaluating their hemolymph osmolality/ions in relation to the outside media. The majority of hemolymph osmolality is comprised of Na+ and Cl− ions (approximately 80%), with other trace ions such as K+, Mg2+, and Ca2+ being essential to the functionality of various transporters such as Na+/K+-ATPase activity, CA, Na+/K+ exchangers, and Na+/K+/2Cl− cotransporter. Although hemolymph free amino acids (FAA) are organic osmolytes that play a role in maintaining osmotic pressure, being particularly important at higher salinities, the contribution is substantially less (Romano and Zeng, ).
Many of the main mechanisms/transporters involved with ion transport are also involved with ammonia-N excretion. Therefore ammonia-N induced osmoregulatory disruptions will only be presented here while the underlying mechanisms explaining these will be discussed at length in the "Uptake Elimination, and Detoxification Mechanisms: Ammonia" section.
It has been reported that ammonia-N reduced hemolymph Cl− in burrowing crab, Neohelice (Chasmagnathus) granulata (Rebelo et al., 1999) and hemolymph Na+ levels in H. americanus (Young-Lai et al., ), M. japonicus (Chen and Chen, 1996), P. leniusculus (Harris et al., ) and S. serrata (Romano and Zeng, ). With the exception of S. serrata, reduced hemolymph Na+ was associated with reduced hemolymph osmolality (Young-Lai et al., ; Chen and Chen, 1996; Harris et al., ), which may indicate that osmotic pressure in S. serrata was maintained via increasing hemolymph FAA, as observed with various crustacean species (Wright et al., ; Durand et al., ; Chen and Chen, ). Interestingly, with the weak osmoregulator, P. pelagicus, ammonia-N levels even approaching lethality had no significant effect on their hemolymph osmolality, Na+, K+, and Ca2+ levels, which may be due to gill Na+/K+-ATPase activity still functioning (Romano and Zeng, ). It should be noted that while osmoregulatory disruptions are sometimes implicated as significant contributors to mortality, it is unlikely a direct cause but rather a consequence/indicator of active ammonia-N excretion becoming disrupted.
It is well known that nitrite-N and Cl− (and possibly Br−) compete for each other during active Cl− uptake (Jensen, ). As previously mentioned, this has toxicity implications, but osmoregulatory consequences as well. For example, nitrite-N levels ≥14 mg L−1 resulted in significantly lower hemolymph Cl− and Na+ levels for P. leniusculus (Harris and Coley, ); significantly lower hemolymph Na+, Cl− and osmolality for M. japonicus (Cheng and Chen, ); significantly lower hemolymph osmolality and Cl− levels for A. astacus (Jensen, ); and significantly lower hemolymph osmolality for P. monodon (Chen and Cheng, ). In the case of A. astacus and M. japonicus, hemolymph FAA significantly increased and was suggested to partially compensate for hemolymph osmolality/ion decreases (Jensen, ; Chen and Cheng, ). However, external Cl− concentrations have been shown to ameliorate the osmoregulatory disruptions during elevated nitrite-N exposure. For M. rosenbergii, an exposure to 10 mg L−1 for 24 hours significantly reduced hemolymph osmolality as well as hemolymph Na+, Cl−, K+, Ca2+, and Mg2+, although these disruptions were not observed when external Cl− levels of the water were increased from 15 to 50 mg L−1 when the same nitrite-N concentration was used (Chen and Lee, ).
Possible reasons for disrupted osmoregulation include increased water uptake thus diluting hemolymph Cl−, Na+, and osmolality, as observed with M. japonicus (Cheng and Chen, ) and/or disruptions to HCO3−/Cl− exchangers (Jensen et al., ). In the case of the latter, elevated nitrite-N levels resulted in significantly lower and higher hemolymph Cl− and HCO3−, respectively, for P. leniusculus (Harris and Coley, ). In a study with A. astacus, Jensen et al. () set out to provide additional evidence to the mechanisms of nitrite-N uptake. Since it is well known that the HCO3−/Cl− exchanger not only functions for osmoregulation but also for acid/base balance, it was hypothesized that hypercapnic conditions (increased CO2 concentrations) would decrease nitrite-N uptake. The reason for this is that some fish compensate for hemolymph acidosis by deliberately inhibiting Cl− uptake that, in turn, would increase HCO3− levels within the blood thus increasing blood pH (Larsen and Jensen, ). It was found that, indeed, both hypercapnic or combined hypercapnic/nitrite-N conditions led to elevated hemolymph HCO3− levels in A. astacus, while hypercapnic conditions reduced nitrite-N uptake, compared to crayfish under normocapnic conditions (Jensen et al., ) thus supporting the suggestion that nitrite-N uptake is via the HCO3−/Cl− exchanger. On the other hand, hypo-osmoregulating animals subjected to higher salinities (those extruding excessive ion influx from the environment) are believed to actively remove nitrite-N back to the environment, possibly via Na+/K+/2Cl− (Jensen, ) and the implications of this will be discussed in the "Uptake, Elimination, and Detoxification Mechanisms: Nitrite" section. Again, studies investigating CA activity on nitrite-N exposed crustaceans may provide additional insight regarding these mechanisms.
Crustaceans possess a hard, calcified exoskeleton that is relatively impermeable to ions (Péqueux, 1995). Therefore, perhaps the most susceptible organ of crustaceans is the gills, since these delicate structures are constantly exposed to external medium and, moreover, perform many crucial physiological processes that include osmoregulation, acid/base balance, gas exchange and ammonia-N excretion (Weihrauch et al., ; Rebelo et al., ; Romano and Zeng, ). However, a limited amount of studies have been performed regarding histopathological changes on decapod crustaceans, despite the essential roles these structures play (Rebelo et al., ; Kuhn et al., ; Romano and Zeng, ).
Exposure to sublethal ammonia-N levels have been shown to cause various gill morphological and physiological changes, including epithelial thickening/sloughing, lamellae constriction/collapse, disruption/destruction to pillar cells, necrosis, and haemocyte infilrations (Rebelo et al., ; Romano and Zeng, ). Due to a nearly complete collapse of gill lamellae when N. granulata were exposed to ammonia-N levels approaching lethality, Rebelo et al. (2000) suggested disrupted gas exchange was one of the main causes for death in this crustacean species. Furthermore, Romano and Zeng () correlated ammonia-N induced gill damage with ammonia-N toxicity through juvenile development of P. pelagicus.
Fortunately, despite ammonia-N causing significant gill damage, it has been demonstrated that, given sufficient time in pristine water, the gills can actually heal again when returned to pristine water (Romano and Zeng, ) (Figure 3). In an experiment with P. pelagicus juveniles, Romano and Zeng () examined the progressive histological gill damage of these crabs subjected to two sublethal ammonia-N concentrations of 10 and 40 mg L−1 for 48 hours and then returned to pristine seawater for 96 hours, and then similarly measured. The results showed that within hours of transfer from elevated ammonia-N exposure to pristine water, the gills gradually healed until near complete healing by 96 hours and was dependent on the recovery time and previous ammonia-N concentration (Romano and Zeng, ) (Figure 3). This may have been facilitated by an increase in THC as well as haemocytes within the gill lamellae (Romano and Zeng, ), since these cells are responsible for phagocytosis of foreign material (Johansson et al., ). However, since such a THC response to ammonia-N exposure is atypical for crustaceans, it is therefore unclear whether a similar result of gill healing would similarly apply to other crustacean species.
Graph: Figure 3 Histopathological changes of the anterior gill lamellae of Portunus pelagicus juveniles previously exposed to 40 mg L−1 ammonia-N and transferred to pristine seawater (no added ammonia-N) at (A) 1 hour, (B) 6 hours, (C) 12 hours, (D) 24 hours, (E) 36 hours, and (F) 48 hours. Note lamellae collapse was prevalent from previous exposure to ammonia-N, while some localized expansions (*) began to occur at 12 hours of recovery. Greater expansions to the entire gill lamellae was observed at increasing recovery durations until a nearly complete recovery by 48 hours including the prevalence of intact pillar cells (PC). Magnification × 20; bars 30 μm.
To the best of our knowledge, only one experiment has been conducted regarding elevated nitrite-N exposure on the histopathological changes to a crustacean, and this was performed on P. pelagicus (Romano and Zeng, ). It was shown that the lowest nitrite-N concentration tested of 3 mg L−1 after 20 days significantly increased gill damage which included increased epithelial lifting, disrupted pillar cells, epithelial thickening and swelling while 6 mg L−1 significantly increased the presence of haemocytes within the gill lamellae (Romano and Zeng, ). Similarly to the ammonia-N induced gill damage, it was suggested that haemocyte infiltration was a response to potentially remove damaged/necrotic tissue and research may be warranted to determine whether gill healing is possible or the role such damage may play with other physiological processes.
It has been shown that nitrate-N can induce significant histopathological changes, although much higher concentrations are required. For P. pelagicus juveniles exposed to the lowest tested nitrate-N level of 1,000 mg L−1 after 4 days, compared to the control, the gills showed drastic and significant changes, including swelling of the lamellae, epithelial thickening, disrupted pillar cells, and necrosis (Romano and Zeng, ). Furthermore, Kuhn et al. () reported that L. vannamei exposed to 35 to 910 mg L−1 nitrate-N over six weeks showed gill "fouling," including excessive debris and bacteria on the gills, while the hepatopancreas exhibited signs of reduced lipid storage and responses to tissue damage that was suggested as a contributor to lower survival and growth. Since these investigations showed significant responses, research should continue in this area as well as possible interactions with other pollutants.
Table 5 Hemolymph ammonia-N levels (mM) of various decapod crustaceans at different concentrations, durations and environmental conditions
|Species||Salinity||and pH||and duration||ammonia-N||Reference|
|Nephrops norvegicus||34||12°C at 7.6–7.8||0 mM||111.4 μM||Schmitt and Uglow (1997)|
|2 mM (10 min)||806.9 μM|
|2 mM (2 hours)||1180.0 μM|
|2 mM (6 hours)||810.0 μM|
|2 mM (9 hours)||606.4 μM|
|4 mM (9 hours)||1456.2 μM|
|Pacifastacus leniusculus||AFW||15°C at 8.2||0 mM||0.30 mM||Harris et al. (2001)|
|15°C at 6.5||1.50 mM (24 hours)||1.49 mM|
|15°C at 8.2||1.50 mM (24 hours)||1.43 mM|
|15°C at 10.5||1.50 mM (24 hours)||0.93 mM|
|Peneaus monodon||34||25°C at 8.2||0 mM (0–10 h)||0.51 mM||Chen and Kou (1993)|
|25°C at 6.3||3.57 mM (8 hours)||0.77 mM|
|25°C at 7.2||3.57 mM (8 hours)||1.24 mM|
|25°C at 8.2||3.57 mM (6 hours)||1.45 mM|
|25°C at 9.0||3.57 mM (1.5 hours)||1.54 mM|
|Eriocheir sinensis||FW||20°C at 7.4||0 mM (3–48 hours)||9.6–8.32 μM||Hong et al. (2007)|
|1.43–5.71 mM (3 hours)||18.95–45.85 μM|
|1.43–5.71 mM (6 hours)||19.80–65.73 μM|
|1.43–5.71 mM (24 hours)||24.35–48.90 μM|
|1.43–5.71 mM (48 hours)||22.84–53.75 μM|
|Scylla serrata||30||28°C at 8.1||0.71–7.14 mM (96 hours)||0.49–5.28 mM||Romano and Zeng (2007b)|
|Portunus pelagicus||15||28°C at 8.1||0 mM (96 hours)||2.0 μM||Romano and Zeng (2010)|
|30||0 mM (96 hours)||0.7 μM|
|45||0 mM (96 hours)||0.7 μM|
|15||1.42–4.28 mM (96 hours)||1.16–4.19 mM|
|30||1.42–4.28 mM (96 hours)||0.79–3.66 mM|
|45||1.42–4.28 mM (96 hours)||0.73–3.39 mM|
|Metacarcinus magister||32||16°C at 8.2||1 mM (1 hours)||326.7 μM||Martin et al. (2011)|
|1 mM (3 hours)||497.0 μM|
|1 mM (7 days)||876.1 μM|
|1mM (14 days)||895.1 μM|
|AFW: Artificial freshwater (NaCl added to be consistent with NH4Cl additions). FW: Freshwater.|
Ultimately, the toxicity and physiological condition of aquatic animals during elevated nitrogenous waste exposure is greatly dependent on its hemolymph accumulation. This is influenced by a variety of factors including gill permeability to the nitrogenous waste species, mechanisms to reduce their passive diffusion, ability to excrete these across a gradient, internal detoxification processes, and environmental factors that alter these adaptive responses.
Three main mechanisms are known for crustaceans to maintain lower hemolymph ammonia-N levels with that of the environment and include active ammonia-N excretion, detoxification and changes/sustaining low gill permeability. It is believed that all three processes can be mobilized simultaneously, although the degree is highly species-specific.
Based on studies thus far, crustaceans have a relatively strong ability to maintain lower hemolymph ammonia-N levels than the outside media (Chen and Kou, ; Schmitt and Uglow, ; Harris et al., ; Hong et al., ; Romano and Zeng, ; ,; Martin et al., ) (Table 5). However, similar factors that influence ammonia-N toxicity (e.g., pH and salinity), also significantly affect hemolymph ammonia-N concentrations. For P. monodon exposed to elevated ammonia-N levels, Chen and Kou () demonstrated that increased pH elevates hemolymph ammonia-N, likely the result of increasing the more permeable NH3 form to passively diffuse across the gills and into the hemolymph. Meanwhile, for P. pelagicus subjected to lower salinities and elevated ammonia-N, hemolymph ammonia-N significantly increased compared to those at higher salinities at the same ammonia-N concentration (Romano and Zeng, ). However, since salinity has less of an influence over NH3 ratios, it was suggested that salinity induced ammonia-N excretion, gill permeability or detoxification alterations were potential causes (Romano and Zeng, ) and these mechanisms will be discussed below.
Generally, crustaceans are well adapted to elevated ammonia-N exposure due to an ecological requirement described by Weihrauch et al. (). Since the majority of decapod crustaceans is benthic and/or exhibit burying behaviors, this increases the likelihood of encountering elevated ammonia-N levels due to their continual production of metabolic by-products as well as sediments generally having higher ammonia-N levels than water-borne levels. Consequently, this was suggested to necessitate the development of active ammonia-N excretion against an inward gradient to prevent excessive hemolymph ammonia-N buildup, thus mitigating their toxic effects (Weihrauch et al., ).
According to this process, which is closely linked with osmoregulation, as the more permeable NH3 diffuses across the lipid bilayers of the gills, it is then protonated to NH4+. NH4+ substitutes for K+ during basolaterally located Na+/K+-ATPase activity, which transports this to the apical membrane on the gills. Finally, NH4+ is excreted to the environment via the apically located transporter Na+/NH4+. In the case of crabs, Weihrauch et al. () proposed the utilization of a functional microtubule network, where NH3 is trapped inside specialized vesicles and is protonated to NH4+ via H+-ATPase. Microtubules then move these to the apical membrane for subsequent exocytosis through the cuticle and into the environment (Weihrauch et al., ). It should be noted that despite the posterior and anterior gills being specialized for osmoregulation and gas exchange, respectively, both are involved with ammonia-N excretion in crustaceans (Weihrauch et al., ; Martin et al., ).
At elevated ammonia-N levels, both ammonia-N excretion and gill Na+/K+-ATPase activity have been shown to increase in various crustacean species, including P. chinensis (Chen and Nan, ), M. nipponense (Wang et al., ) and P. pelagicus juveniles (Romano and Zeng, 2010, ). Further, in an in vitro experiment, the presence of ammonia-N in test tubes increased gill Na+/K+-ATPase activity for Macrobrachium olfersii (Furriel et al., 2004), Callinectes danae (Masui et al., 2002, 2005) and C. ornatus (Garçon et al., 2007). Species-specific osmoregulatory abilities was suggested to greatly influence ammonia-N excretion rates (Weihrauch et al., ; Martin et al., ). In an experiment with three different crabs, Weihrauch et al. () acclimated E. sinensis, Carcinus maenas, and Cancer pagurus, representing strong, moderate, and weak osmoregulators, to salinities of 0, 10, and 35, respectively, and measured their ammonia-N excretion rates. It was revealed that C. pagurus exhibited the highest rates of ammonia-N excretion against an eightfold inward gradient, followed by C. maenas and E. sinensis (capable of excreting ammonia-N against approximately a fourfold inward gradient) (Weihrauch et al., ). This indicated that since E. sinensis possessed relatively low gill permeabilities to ammonia-N, likely linked with their strong osmoregulatory abilities, high rates of ammonia-N excretion were not required (Weihrauch et al., ). More recently, Martin et al. () demonstrated that for Dungeness crab, Metacarcinus magister, this weak osmoregulator was capable of excreting ammonia-N against a 16-fold inward gradient.
However, as previously stated, there is a point when this mechanism begins to break down. For M. magister, it was observed that mRNA expression of many transporters responsible for ammonia-N excretion, including Na+/K+-ATPase, H+-ATPase, Na+/H+ transporter and Rhesus-like protein transporter (RhMM), were down-regulated prior to or at 14 days of exposure to 14 mg L−1 ammonia-N and coincided with a cessation to ammonia-N excretion (Martin et al., ). Martin et al. () noted that since none of the crabs experienced any outwards signs of stress this pointed to additional coping mechanisms which may have included an ability to reduce passive NH3 inward diffusion (by closing certain ion channels on the gills) and/or detoxification mechanisms.
In the case of the latter, ammonia-N detoxification has been reported in numerous crustacean species (Chen and Cheng, ; Chen and Chia, 1996; Durand et al., ; Chen and Chen, ; Lee and Chen, 2003; Martin et al., ) and is an important adaptation since hemolymph accumulation is inevitable during their prolonged and/or elevated exposure. Using emersed crabs (i.e., when out of water), which all but precludes the possibility of active ammonia-N excretion, it was found that C. maenas significantly increased non-essential amino acids (NEAA) levels within their muscle (Durand et al., ). The experiment was set up to exclude hyper-osmotic stress, protein degradation and dietary origin factors as potential causes/origins for NEAA production, and therefore, Durand et al. () suggested such an increase was the use of ammonia-N derived nitrogen as a source for de novo NEAA synthesis. A similar mechanism may also be used by the isopod, Porcellio scaber, since exposure to high gaseous ammonia also led to a significant increase in NEAA (Wright et al., ). Similarly, for P. monodon juveniles exposed to elevated water-borne ammonia-N, this led to significantly increased both essential and non-essential hemolymph FAA while E. sinensis increased hemolymph glutamine (Chen and Chen, ; Hong et al., ). In addition to the suggestion of de novo synthesis, it may have also been an adaptive response to maintain osmotic pressure when hemolymph Na+ levels decrease (Chen and Chen, ). However, in the case of subtidal velvet crab, Necora puber, Durand et al. () found no significant increase to FAA during emersion, although it was suggested that species may have utilized different ammonia-N detoxification methods that were not measured.
In addition to hemolymph FAA production, ammonia-N is believed to be detoxified to urea-N and/or nitrite-N in some species. In the case of latter, there was a significant increase to nitrite-N excretion when P. monodon were exposed to 4.63 mg L−1 ammonia-N (Chen and Cheng, ). After measuring the test solutions to confirm that nitrite-N was absent, Chen and Cheng () suggested hemolymph ammonia-N was converted to the less toxic nitrite-N and can subsequently be passively diffused outwards. In the same experiment, Chen and Cheng () also measured significantly increased urea-N excretion, along with a concomitant decrease to ammonia-N excretion, when P. monodon were exposed to increasing ammonia-N levels, thus likely indicating ammonia-N detoxification to urea-N. Similar processes have also been demonstrated in other crustaceans including S. serrata (Chen and Chia, 1996), C. maenas (Durand et al., ), P. leniusculus (Harris et al., ), M. japonicus (Lee and Chen, 2003), E. sinensis (Hong et al., ), and M. magister (Martin et al., ). Interestingly, higher salinities or pH were shown to increase urea-N production in S. serrata and M. japonicus (Chen and Chia, 1996; Lee and Chen, 2003) and P. leniusculus (Harris et al., ), respectively. Durand et al. () noted that this process is likely a highly efficient detoxification method since 1 μmol of urate requires 4 μmol of amino acid precursors (i.e., nitrogen).
Table 6 The hemolymph nitrite-N levels (mM) of various crustaceans following different exposure concentrations, durations, and environmental conditions
|Species||Salinity or Cl−||and pH||and duration||concentration||Reference|
| Procambrus clarkii||FW (22 mg L−1 Cl−)||23°C at 7.9–8.3||5.71 mM (24 hours)||≈42.83 mMA||Gutzmer and Tomasso (1985)|
|FW (100 mg L−1 Cl−)||23°C at 7.9–8.3||5.71 mM (24 hours)||≈62.81 mM|
| Pacifastacus leniusculus||FW (44.6 mg L−1 Cl−)||7.6||1.0 (24 hours)||25.00 mM||Harris and Coley (1991)|
| Astacus astacus||FW (46.1 mg L−1 Cl−)||15°C at 8.3||1.0 (7 days)||17.60 mM||Jensen (1996)|
| A. astacus||FW (46.4 mg L−1 Cl−)||6°C at 8.3||1.0 (48 hours)||5.00 mM||Jeberg and Jensen (1994)|
|16°C at 8.3||10.00 mM|
| A. leptodactylus||FW (35 mg L−1 Cl−)||13°C at 7.21||1.78 mM (48 hours)||61.38 mM||Yildiz and Benli (2004)|
|FW (100 mg L−1 Cl−)||1.71 mM (48 hours)||2.50 mM|
| Macrobrachium rosenbergii||FW (15 mg L−1 Cl−)||27°C at 7.83||0.713 mM (24 hours)||2.16 μM||Chen and Lee (1997)|
|FW (33 mg L−1 Cl−)||0.67 μM|
|FW (50 mg L−1 Cl−)||0.01 μM|
| Peneaus monodon||25||25°C at 6.8||0.72 mM (48 hours)||3.70 μM||Chen and Cheng (2000)|
|25°C at 8.2||2.87 μM|
|25°C at 9.8||1.91 μM|
| P. japonicus||25||20°C||1.43 mM (16 hours)||1.70 μM||Chen and Chen (1992b)|
| P. japonicus||30||25.3°C at 8.28||1.43 mM (24 hours)||2.77 μM||Cheng and Chen (2001)|
|1.43 mM (48 hours)||4.48 μM|
|AData only presented in figures and approximate values are presented.|
Nitrite-N concentrations within the hemolymph of freshwater crustaceans or those subjected to lower salinities often far exceed that of the external media (Table 6), which likely explains higher nitrite-N toxicities (Jensen, ; Lin and Chen, 2003). For freshwater crayfish, A. astacus, exposed to 11.2 mg L−1 (0.8 mM) nitrite-N in freshwater, a more than tenfold increase to hemolymph nitrite-N was detected of 140 mg L−1 (10 mM) within 48 hours (Jensen, ), while in another experiment on the same species, the hemolymph nitrite-N level was more than 17-fold higher at 246.6 mg L−1 (17.6 mM) when subjected to 14.0 mg L−1 (1 mM) nitrite-N after 7 days (Jensen, ). An even greater discrepancy was measured for freshwater crayfish, P. leniusculus, of 350.3 mg L−1 (25 mM) hemolymph nitrite-N when exposed to 14.0 mg L−1 (1 mM) nitrite for 24 hours (Harris and Coley, ).
Based on the various protective effects of Cl− to nitrite-N exposure mentioned previously, it is not surprising that Cl− can significantly inhibit hemolymph nitrite-N uptake in various crustacean species (Chen and Chen, ,; Chen and Lee, ; Yildiz and Benli, ). For example, when Cl− concentrations were 15 mg L−1 and 50 mg L−1, hemolymph nitrite-N levels of M. rosenbergii were 30.3 μg L−1 and 0.13 μg L−1 (2.16 μM and 0.01 μM), respectively when subjected to 10 mg L−1 (0.71 mM) nitrite-N for 24 hours (Chen and Lee, ). Moreover, after 48 hours, the hemolymph nitrite-N levels were more than 24-fold lower in A. leptodactylus when 100 mg L−1 Cl− was added, compared to tap water conditions (33 mg L−1 Cl−) at similar nitrite-N levels (Yildiz and Benli, ). Meanwhile, for P. monodon and M. japonicus exposed to an mM nitrite range, the hemolymph nitrite-N levels were in a μM range when higher salinities between 25–30 were employed (Chen and Chen, ,; Chen and Cheng, ). The cause for this is that hypo-osmoregulating crustaceans would actively extrude Cl−/nitrite-N to the environment, while the opposite is the case of hyper-osmoregulating animals (Jensen, ). Temperature has also been shown to a significant factor, albeit to a less extent (Jeberg and Jensen, ) (Table 6).
Similarly to elevated ammonia-N exposure, there is evidence indicating an ability of crustaceans to detoxify nitrite-N (Chen and Cheng, ; Jensen, ; Cheng and Chen, 1998, 2001). It has been reported that increasing nitrite-N levels (from 0.07 to 1.43 mg L−1) and duration of exposure leads to higher hemolymph urea-N concentrations in M. japonicus (Chen and Cheng, ; Cheng and Chen, 1998, 2001) and P. monodon (S. Y. Cheng and Chen, ). Two possible explanations have been suggested. The first suggests that urea may be created through either the hydrolysis of arginine or during the ornithine-urea cycle since both ammonia-N and arginine also decreased within the hemolymph of M. japonicus (Cheng and Chen, 2001). The second explanation, proposed by Cheng and Chen (2001), is that since hemolymph HCO3− decreased (along with ammonia-N), may indicate urea-N synthesis when ammonia-N and HCO3− react with one another as reported with toadfish, Opsanus beta (Walsh et al., ). Moreover, Jensen () reported that more than 50% of nitrite-N in A. astacus was likely converted to nitrate-N evidenced by elevated hemolymph nitrate-N levels in the absence of nitrate-N levels in the test solution.
The processes of nitrate-N uptake and elimination in crustaceans is substantially less complex since active transport does not appear to be involved, while gill permeability to nitrate-N is likely low, thereby reducing passive diffusion (Jensen, ). In an experiment comparing nitrite-N and nitrate-N uptake in A. astacus, it was shown that nitrate-N concentrations within the hemolymph were more than 15-fold less than the outside media (approximately 0.06 mM nitrate-N within the hemolymph versus 1 mM nitrate-N in the test solutions) after seven days (Jensen, 1996). Similarly, Cheng et al. (2002) measured nitrate-N accumulation within various organs of P. monodon when exposed to 3.64 mM nitrate-N after 24 hours, with the highest and lowest nitrate-N concentrations detected within the midgut (1.53 mM nitrate-N) and muscle (0.02 μM nitrate), respectively. Consequently, it is believed that one of the main reasons nitrate-N is less toxic is due to reduced hemolymph accumulation (Jensen, ).
Excessive nitrogen loading in ecosystems associated with climate change and increased anthropogenic activities are increasingly becoming an issue; however, the implications of nitrogenous waste exposure to crustaceans have often been overlooked due to a perceived low toxicity. Indeed, although decapod crustaceans appear particularly well adapted to these pollutants, the traditional use of the "safe" concentrations, obtained from 96-hour LC50 values (Sprague, ), is likely a significant underestimation and a misleading value in relation to nitrogenous waste. In addition to being an inaccurate prediction of growth, various physiological processes become disrupted at or substantially below these values, which can include osmoregulation, immunology, acid/base balance, and gas exchange as well as increasing oxidative stress, pathogenic susceptibility, and histopathological damage. While this should not de-emphasize the importance of measuring LC50 values since important patterns can be obtained, physiological data can provide a valuable compliment. In particular, to our knowledge no information currently exists regarding the effects of these nutrients on crustacean reproduction, such as egg viability, hatching rates, or lipid deposits, and appears to be an area worth investigating. Moreover, currently the majority of research generally focuses on one pollutant to crustaceans; however, this may not be representative of real life scenarios in either nature or an aquaculture setting and therefore determining potential interactions with nitrogenous wastes or other pollutants may also be explored. Such research directions are important to help elucidate causes for reduced abundance and/or physiological condition of crustaceans in various ecosystems receiving excessive effluent discharges as well as under cultured conditions since these will likely have significant implications to the multi-billion dollar fisheries and aquaculture industry. REFERENCES
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By Nicholas Romano and Chaoshu Zeng
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