Review
Acclimation and toxicity of high ammonium concentrations to unicellular algae

https://doi.org/10.1016/j.marpolbul.2014.01.006Get rights and content

Highlights

  • Growth rates of 6 microalgal classes in cultures were compiled along with ammonium concentrations.

  • Chlorophyceae and cyanophyceae were the most tolerant to high ammonium levels.

  • Diatoms do not seem to be seriously affected by ammonium concentrations below 100 μM.

  • Dinoflagellates appear to be least tolerant to high ammonium levels.

  • Field ammonium concentrations <100 μM will not likely reduce the growth of most microalgae.

Abstract

A literature review on the effects of high ammonium concentrations on the growth of 6 classes of microalgae suggests the following rankings. Mean optimal ammonium concentrations were 7600, 2500, 1400, 340, 260, 100 μM for Chlorophyceae, Cyanophyceae, Prymnesiophyceae, Diatomophyceae, Raphidophyceae, and Dinophyceae respectively and their tolerance to high toxic ammonium levels was 39,000, 13,000, 2300, 3600, 2500, 1200 μM respectively. Field ammonium concentrations <100 μM would not likely reduce the growth rate of most microalgae. Chlorophytes were significantly more tolerant to high ammonium than diatoms, prymnesiophytes, dinoflagellates, and raphidophytes. Cyanophytes were significantly more tolerant than dinoflagellates which were the least tolerant. A smaller but more complete data set was used to estimate ammonium EC50 values, and the ranking was: Chlorophyceae > Cyanophyceae, Dinophyceae, Diatomophyceae, and Raphidophyceae. Ammonia toxicity is mainly attributed to NH3 at pHs >9 and at pHs <8, toxicity is likely associated with the ammonium ion rather than ammonia.

Introduction

Nitrogen (N) is the element that generally limits phytoplankton growth in many coastal waters and oceans (Ryther and Dunstan, 1971, Boynton et al., 1982, Hecky and Kilham, 1988, Howarth and Marino, 2006). The forms of inorganic N have been suggested to structure phytoplankton communities in a variety of environments, i.e. nitrate leading to diatom blooms (Lomas and Glibert, 1999) and ammonium stimulating flagellates/dinoflagellates growth (Boynton et al., 1982, Malone et al., 1983, Price et al., 1985, Probyn, 1985, Robert et al., 1986, Semeneh et al., 1998). In addition to the form of N, high concentrations of ammonium that often occur near marine outfalls have also been shown to influence phytoplankton blooms. For example, Keller et al. (1987, see their Table 2) reported that 100 μM ammonium was toxic for 200 species or clones of oceanic phytoplankton as determined by fluorescence. Recently, high ammonium was suggested to be detrimental to normal coastal diatom development in urbanized estuaries (Wilkerson et al., 2006, Dugdale et al., 2007, Yoshiyama and Sharp, 2006, Glibert, 2010, Parker et al., 2012a, Parker et al., 2012b, but see Cloern et al., 2012). More specifically, it was argued that ammonium inhibited nitrate uptake and delayed diatom blooms (Dugdale et al., 2007), depressed primary production and suppressed diatom spring blooms (Parker et al., 2012a), or led to low assimilation numbers (Yoshiyama and Sharp, 2006). In contrast, ammonium has been reported to stimulate diatom growth relative to nitrate not only in cultures (Harvey, 1953, Thompson et al., 1989, Livingston et al., 2002, Bender et al., 2012), but also in the field (Harris, 1959, Takahashi and Fukazawa, 1982, Suksomjit et al., 2009b, Tada et al., 2009).

In order to try to reconcile these opposing lines of evidence, the literature on the effects of ammonium concentrations on marine phytoplankton growth in laboratory cultures is reviewed. Optimal, inhibitory and toxic concentrations were identified among the different classes of unicellular algae and the mechanisms responsible for ammonium toxicity are briefly discussed. Some acclimation mechanisms such as lag phases and multi-phasic uptake systems were also identified and their consequences are discussed in the present context of eutrophication in coastal zones. Since ammonium toxicity is also a problem in sewage oxidation ponds, and pulp mill effluents, the removal of ammonium is necessary in wastewater treatment and in high capacity fish/shrimp ponds. These examples are briefly discussed as well.

Ammonium toxicity in water can be due to the effects of both the unionized ammonia (NH3) and the ionized ammonium (NH4+). Ammonia is considered to be the most toxic form because it is uncharged and lipid soluble and easily diffuses across membranes. Since it is a gas, it is volatile and can be lost to the atmosphere, especially in actively aerated cultures. In contrast to ammonia, the charge on the membrane hinders the passage of the charged ammonium ion. There is no chemical method that can measure these two forms separately. Present chemical methods measure both forms that are often termed ‘total ammonia’ (i.e. NH4 + NH3). The relative concentration of each form is strongly dependent on pH and to a lesser extent on temperature, and salinity has only a minor influence (Whitfield, 1974, Emerson et al., 1975, Bower and Bidwell, 1978, Spotte and Adams, 1983). In general, at the pH of seawater at 8.0 and 20oC, only about 10% of the total ammonia is present as the more toxic form, ammonia. Since 90% is present as the ammonium ion, it is preferable to use the term ammonium for natural seawater. As pH increases, the concentration of ammonia increases dramatically (Fig. 1). For example, in freshwater fish ponds at 30oC, when the pH increases from 7.0 to 9.0, the ammonia concentration increases over 60-fold, but it is still only 45% of the ammonia + ammonium concentration. In this case, the term ammonia (actually total ammonia as determined by the chemical method) is generally used in freshwater aquaculture since the concern is on ammonia toxicity to fish. Emerson et al. (1975; see their Table 2) and Spotte and Adams (1983; see their Table 1) give the %NH3 for a range of pHs and temperatures. The ratio of unionized ammonia to ammonium ion increases by 10-fold for each unit increase in pH and only 2-fold for each 10 °C rise in temperature over the 0–30 °C range (Erickson 1985). Bower and Bidwell (1978) also included the influence of salinity as well as pH and temperature in their four tables. An increase in the ionic strength of the solution (i.e. an increase in salinity or water hardness in freshwater) causes only a very small decrease in the %NH3. For example, an increase in salinity from 20 to >34 causes a small decrease in %NH3 from 3.41 to 2.98 (Bower and Bidwell, 1978). The dissociation constant (pKa) of the ammonia/ammonium reaction is about 9.3 depending a salinity, temperature, etc. In summary, ammonia toxicity is almost solely attributed to NH3 at higher pHs of about 9 and at pHs <8, any toxicity effects are more likely associated with the ammonium ion than ammonia (Erickson, 1985). At a pH between 8 and 9, Erickson (1985) found that toxicity to fish was due to both ammonia and ammonium and similarly for other organisms (USEPA, 1999). In seawater at 20 °C, pH 8 and a salinity of 30, ammonia comprises 3.8% of the total ammonia (ANZECC/ARMCANZ, 2000), with ammonium contributing <1% of the total toxicity (USEPA, 1989, Batley and Simpson, 2009). For consistency in this review, we use ammonium, even though both ammonia and ammonium are present at most pHs discussed here, and we have used the units of μM and converted mg L−1, the unit that is used in freshwater studies, to μM.

During growth of algae on ammonium, the pH decreases due to the release of H+ ions to the medium. In contrast, growth on nitrate causes an increase in the pH due to the release of OH- ions and during growth on urea, there is little change in the pH (Raven and Smith, 1976, Goldman et al., 1982a, Goldman et al., 1982b, Goldman et al., 1982c, Raven, 1988, Britto et al., 2001b).

Generally, ammonium concentrations in surface waters range up to ∼3 μM. In contrast, anthropogenic inputs of ammonium from atmospheric deposition, agricultural activities and sewage are regarded as ‘new’ nitrogen (Dugdale and Goering, 1967). Examples of anthropogenically produced atmospheric sources are transportation emissions, and volatilization from manure produced in animal farming. Atmospheric deposition can range from 10 to over 40% of the new N loading to estuaries that are downwind of anthropogenic emissions (Paerl et al., 2002). On the east coast of the USA, atmospheric deposition can account for 10–40% of the new nitrogen loading to estuaries and may exceed riverine input in many areas. The air shed may exceed the watershed by 10–20-fold (Paerl et al., 2002). Agricultural activities, including an increase in intensive animal farming (especially pigs and chickens) and the liberal use of fertilizer, have increased ammonium loading to the coastal zone via atmospheric wet and dry deposition and groundwater (Raven et al., 1992).

In the N cycle in marine ecosystems, ammonium is usually regarded as ‘regenerated’ N and is produced by excretion from animals and by bacterial regeneration/recycling of organic N compounds in the sediments and water column (Galloway et al., 2003, Galloway et al., 2004). Another minor source of ammonium in sediments that is produced via nitrogen cycle processes is dissimilatory nitrate reduction (DNR). In sediments with high sulphide concentrations, nitrification and denitrification may be inhibited, but dissimilatory nitrate reduction to ammonium may be enhanced because sulfide acts an electron donor (An and Gardner, 2002, Brandes et al., 2007). Enhanced DNR and reduced denitrification may preserve available nitrogen in estuarine sediments.

In anoxic, high organic sediments, the N transformation reactions stop at the conversion of organic N to ammonium. Since ammonium is highly soluble, it is recycled via pore water to the water column. The primary sinks for ammonium include phytoplankton and possibly bacterial uptake and microbial nitrification (NH4  NO3). Nitrate may be converted to N2 via denitrification processes when oxygen is low. Under anoxic conditions, anaerobic oxidation of ammonium can occur (NH4 + NO2-N2; called the anammox reaction; Galloway et al., 2003, Brandes et al., 2007).

The highest values of ammonium are generally found in freshwater environments. Maxima ranged from 100 μM in Wascana Lake, Canada (Donald et al., 2011), 200 μM in prairie lakes, Canada (Murphy and Brownlee, 1981), 300 μM in Lake Taihu, China (Chen et al., 2003) and up to 680 μM recorded in summer 1990 in Lake Little Mere, UK (Carvalho, 1994). Maxima sometimes increased over time such as in Lake Donghu, China from 10 μM in the 1950s to 60 μM in the 1980s and up to 100 μM in 2001 (Dai et al., 2012). In lagoons, maxima ranged from 9 μM in Thau lagoon, France (Collos et al., 2005), to 300 μM in Bolmon lagoon, France (Chomérat et al., 2007) and Balearic islands lagoons (Lucena-Moya et al., 2012) and up to 400 μM in Portuguese lagoons (Coutinho et al., 2012).

In estuaries, ammonium concentrations are generally related to salinity (S), with high values in the less saline part, such as 300 μM in the Ems-Dollard estuary (Admiraal, 1977), >400 μM in Deep Bay, Hong Kong (Xu et al., 2010) or 1000 μM in the Colne estuary (UK) at S = 10, decreasing to 100 μM at S = 20 and 20 μM at S = 30 (Underwood and Provot, 2000). In coastal waters, the highest values are found near sewage outfalls: >25 μM in Victoria Harbour, Hong Kong (Xu et al., 2008), 40 μM in Santa Monica Bay, California (MacIsaac et al., 1979), 150 μM at Whites Point, California (Thomas and Carsola, 1980), 3000 μM in the Ems-Dollard estuary (Admiraal, 1977). However, high values can also be found at salinities over 30, such as in Annaba Bay (Algeria) where 100 μM was recorded in summer (Ounissi and Fréhi, 1999).

While temporal increases can be found in some marine environments such as in Osaka Bay, increasing from 30 μM in 1980 (Yamochi and Abe, 1984) to 300 μM in 2007 (Yamamoto et al., 2010), other maxima seem to be more stable such as in Suisun Bay (USA): 27 μM in 1974 (Glibert, 2010), 16 μM in 2000–2003 (Wilkerson et al., 2006), 14 μM in 2006 (Parker et al., 2012a, Parker et al., 2012b). High frequency sampling has revealed large diel changes in ammonium concentrations that could increase 50-fold (from 0.1 to 5 μM) during the night possibly due to grazing (Litaker et al., 1988, Yamamuro and Koike, 1994, Horner-Rosser and Thompson, 2001).

Fig. 2 shows the response of an estuarine diatom to a range of concentrations of nitrate, nitrite and ammonium (Rao and Sridharan, 1980). Notice the log scale and the large difference in the toxic concentration between nitrate and ammonium. It illustrates that all three inorganic N substrates can become inhibitory above a certain concentration. A similar experiment was conducted with ten species of estuarine benthic diatoms and good growth occurred at 17,000 μM nitrate, 1000–10,000 μM nitrite, but only 500 μM ammonium was inhibitory (Rao and Sridharan, 1980). It is interesting to see how few studies have compared the toxicity of these three inorganic N species. Here we focus on ammonia/ammonium toxicity in particular, even though other inorganic N forms can become toxic at high levels. At concentrations that are not toxic, ammonium has frequently been reported to produce higher growth rates compared to nitrate and urea for a wide variety of species (Paasche, 1971, Thompson et al., 1989, Giordano, 1997, Suksomjit et al., 2009a, Tada et al., 2009, Hii et al., 2011).

Overall, laboratory culture data from 45 freshwater and 68 marine studies were used. This review is not intended to be exhaustive, but it should be fairly representative of available data on the effect of ammonium on phytoplankton growth in a wide variety of aquatic environments. Unfortunately, many laboratory studies did not measure pH of the culture medium at the end of the growth period and therefore it was not always possible to estimate the%NH3 in the medium and its potential toxicity. Some studies used pH buffers to keep pH relatively constant during batch culture growth. In addition, many studies did not use ecologically important species. Data were also examined for indications of possible acclimation of unicellular algae to high ammonium levels and the%NH3 at known pHs.

Names of microalgae were checked against the Algaebase (http://algaebase.org/) or the WoRMS (World Register of Marine Species) database and currently accepted names were used whenever possible.

We used three categories of ammonium concentrations (e.g. optimal, inhibitory and toxic) in relation to phytoplankton growth rates that were usually assessed by an increase/decrease in cell counts and/or in vivo fluorescence (i.e. long term effects). In a few studies, growth rate decreased, but not cell yield, or vice versa. Short term effects of ammonium additions based on 14C uptake (Azov and Goldman, 1982, Collos, 1986 and references therein; Turpin, 1991, Huppe and Turpin, 1994) or oxygen production (Belkin and Boussiba, 1991) are not included here. Fluorometric estimates of photosynthetic efficiency are mentioned where useful. “Optimal” concentrations were defined as those leading to maximal growth and were experimentally determined by measuring growth rates over a range of ammonium concentrations. If no gradient was used, optimal concentration data were included if authors indicated that they were optimal (Chu, 1942, Chu, 1943, Guillard and Wangersky, 1958, Kapp et al., 1975, Kim et al., 2012, Pintner and Provasoli, 1963, ZoBell, 1935), or if growth on ammonium was greater than growth on nitrate (i.e. the control) at equimolar concentrations (Ryther, 1954, Stewart, 1964, Moss, 1973, Fabregas et al., 1989, Thakur and Kumar, 1999, Giordano, 2001, Shi et al., 2000, Suksomjit et al., 2009a, Suksomjit et al., 2009b, Chen et al., 2011, Hii et al., 2011). It was often observed that there was a lag phase in growth that increased as the initial ammonium concentration increased in batch cultures (Collos, 1986 and references therein, Bates et al., 1993, Matsuda et al., 1999, Collos et al., 2004, Nagasoe et al., 2010, Park et al., 2010), without affecting the growth rate reached after the acclimation period. We noted this lag in final maximal growth rates where applicable. Inhibitory concentrations significantly reduced growth rate compared to optimum concentrations and were represented by the EC50, (the effective concentration where growth rate was reduced by 50%, a common term used in ecotoxicology studies). The toxic concentration is the concentration at which no growth was observed.

Section snippets

Effect of ammonium on growth rates of unicellular algae

Table 1 summarizes optimum, inhibitory and toxic ammonium concentrations (details in Supplementary tables) for growth rates of six classes of unicellular algae. The ranking of the six algal classes in terms of their tolerance to high ammonium levels was as follows:

Chlorophyceae > Cyanophyceae > Diatomophyceae > Raphidophyceae > Prymnesiophyceae > Dinophyceae.

The Kruskal–Wallis test and Dunn’s multiple comparison test revealed that for toxic concentrations, Chlorophytes were significantly more tolerant to

Acclimation time for ammonium uptake at high ammonium concentrations

In the previous section, we reviewed the effect of ammonium on growth rates which are relatively long term and integrate short term physiological processes. Here we review initial short-term physiological responses of algal cells suddenly exposed to a pulse of ammonium that can occur near effluent sources. These short-term transient responses over a few hours may or may not translate into long term effects on community composition since a brief lag/induction/acclimation period, for example, may

Bioassays and water quality criteria for ammonium

Direct toxicity assessment or whole effluent toxicity testing is an important part of the regulatory framework in many countries and is used for compliance monitoring of various effluents and contaminated waters (USEPA, 1989, USEPA, 1999). Guidelines for toxicity testing recommend that a range of organisms be used, including micro and macroalgae, echinoderms, bivalves, a gastropod, crustaceans and fish. It is important that pH be kept constant during the tests because variations in pH can

Fish ponds/farms

In fish aquaculture, ammonia toxicity is a common problem for the fish. It is a by-product of protein metabolism from their high protein diet and it is excreted from the gills. This production of ammonia from the fish is taken up by phytoplankton, but when the algal blooms crash, ammonia increases and toxicity for the fish may occur. There is interest in finding phytoplankton species that can tolerate high ammonium concentrations and reduce ammonium concentrations for the more sensitive fish.

Conclusions

The effects of ammonium on microalgae can occur on two different time scales: long term growth rates (days) and short term physiological processes such as uptake rates, photosynthetic rates, and enzyme activities that occur over minutes to hours. Growth rates generally integrate these short term transient physiological responses to changes in ammonium concentrations that occur over a few hours. These short term responses may or may not translate into changes in long term community composition

Acknowledgements

We thank Patricia Glibert, Richard Dugdale, Francis Wilkerson, Jim Cloern and an anonymous reviewer for their comments that improved the manuscript.

Y.C. acknowledges support from CNRS.

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