Acclimation and toxicity of high ammonium concentrations to unicellular algae.

Marine Pollution Bulletin 80 (2014) 8–23

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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul

Review

Acclimation and toxicity of high ammonium concentrations to unicellular algae Yves Collos a,⇑, Paul J. Harrison b a b

Ecologie des Systèmes Marins Côtiers (UMR5119), Université Montpellier 2, CNRS, IRD, case 093, 34095 Montpellier Cedex 5, France University of British Columbia, Department of Earth & Ocean Sciences, Vancouver, BC V6T 1Z4, Canada

a r t i c l e

i n f o

Keywords: Ammonia/ammonium Toxicity Phytoplankton Acclimation EC50 for ammonia pH

a b s t r a c t 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 lM 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 lM respectively. Field ammonium concentrations Cyanophyceae, Dinophyceae, Diatomophyceae, and Raphidophyceae. Ammonia toxicity is mainly attributed to NH3 at pHs >9 and at pHs 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 25 lM in Victoria Harbour, Hong Kong (Xu et al., 2008), 40 lM in Santa Monica Bay, California (MacIsaac et al., 1979), 150 lM at Whites Point, California (Thomas and Carsola, 1980), 3000 lM 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 lM 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 lM in 1980 (Yamochi and Abe, 1984) to 300 lM in 2007 (Yamamoto et al., 2010), other maxima seem to be more stable such as in Suisun Bay (USA): 27 lM in 1974 (Glibert, 2010), 16 lM in 2000–2003 (Wilkerson et al., 2006), 14 lM in 2006 (Parker et al., 2012a,b). High frequency sampling has revealed large diel changes in ammonium concentrations that could increase 50-fold (from 0.1 to 5 lM) during the night possibly due to grazing (Litaker et al., 1988; Yamamuro and Koike, 1994; Horner-Rosser and Thompson, 2001). 1.4. Generalized response of growth rate to dissolved inorganic nitrogen concentrations 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 lM nitrate, 1000–10,000 lM nitrite, but only 500 lM 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). 1.5. Sources of data 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

Fig. 2. Growth rate (in units of day1) of the diatom Pleurosigma aestuarii as a function of nitrogen concentrations (data from Rao and Sridharan, 1980). Ammonium (triangles), nitrite (squares), nitrate (diamonds).

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Y. Collos, P.J. Harrison / Marine Pollution Bulletin 80 (2014) 8–23

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, 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,b; 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. 2. 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 > Diatomo phyceae > Raphidophyceae > Prymnesiophyceae > Dinophyceae.

The Kruskal–Wallis test and Dunn’s multiple comparison test revealed that for toxic concentrations, Chlorophytes were significantly more tolerant to high ammonium than diatoms (p < 0.001), dinoflagellates (p < 0.001), and raphidophytes (p < 0.05). Cyanophytes were significantly more tolerant than dinoflagellates (p < 0.01). In general, comparing the six classes, dinoflagellates were the least tolerant to high ammonium levels. The ranking for inhibitory and optimal levels were similar. For the Chrysophyceae, Keller et al. (1987) were able to grow Ochromonas sp. at 100 lM ammonium and Watson and McCauley (2005) found optimal concentrations at 140 lM for Uroglenopsis americana and Synura petersenii and at 14 lM for Dinobryon cylindricum. The toxic level for the latter three species was identical at 1500 lM. For the Cryptophyceae, only three studies were found (Antia and Chorney, 1968; Guillard and Wangersky, 1958; Lourenço et al., 2002), with toxic ammonium levels of 5000, 1200 and 11,000 lM for Hemiselmis virescens, Hillea sp., and Rhodomomas sp. respectively. Concerning the Dictyochophyceae, optimal growth of Aureococcus anophagefferens occurred at 50 lM ammonium, with inhibitory and toxic ammonium levels at 200 and 500 lM respectively (Fan et al., 2003; Taylor et al., 2006). A smaller data set was used that included gradients of ammonium concentrations and their effects on growth rates (Table 2). EC50 values were provided by investigators or could be calculated from raw data. The EC50 for growth ranged from 30 to 56,230 lM. The ranking of EC50 values was slightly different from the one derived from Table 1: Chlorophyceae > Cyanophyceae > Diatomophyceae > Dinophyceae > Raphidophyceae, but the dinoflagellates, the diatoms and the raphidophytes were not significantly different. Only the Chlorophyceae were significantly more tolerant to ammonium than diatoms (p < 0.001), dinoflagellates (p < 0.01) and the Cyanophyceae were more tolerant than the Dinophyceae. When pH and temperature data were provided, we calculated the ammonia concentration (rather than ammonia + ammonium values that were used above) and determined EC50 values for ammonia toxicity for the five algal groups. We found the following ranking: Cyanophyceae > Chlorophyceae > Diatomophyceae > Raphidophyceae > Dinophyceae, but the chlorophytes, the dinoflagellates, the diatoms and the raphidophytes were not significantly different. Only the Cyanophyceae were significantly more tolerant to ammonia than diatoms (p < 0.05) and dinoflagellates (p < 0.01). Many of the effects of ammonia on growth rates were determined in cultures grown at temperatures of 18 °C or higher. As temperature increases from 10 to 30 °C, the %NH3 also increases by 20% at pH 9.0, but only slightly at a pH of 8.0 (Fig. 1). Therefore, growth rates at lower temperatures should be less sensitive to a certain total ammonium concentration than at higher temperatures since NH3 is more toxic than ðNHþ 4 Þ. As discussed above for Table 2, if the temperature and pH of the cultures were given, then we calculated the EC50 values for NH3 rather than total ammonia

Table 1 Optimal, inhibitory and toxic ammonium concentration (lM) for growth of unicellular algae in batch cultures. Optimal concentrations were determined from data for which at least one adverse concentration was identified or when the authors specifically mentioned they were optimal. Number of studies is in brackets. Class

Mean ± SD Optimal

Mean ± SD Inhibitory

Mean ± SD Toxic

Chlorophyceae Cyanophyceae Diatomophyceae Dinophyceae Prymnesiophyceae Raphidophyceae

7572 ± 7619 (17) 2486 ± 1570 (10) 337 ± 409 (27) 110 ± 77 (23) 1432 ± 1197 (12) 263 ± 332 (8)

23,758 ± 25,922 (14) 6616 ± 6514 (10) 725 ± 839 (27) 324 ± 283 (25) 958 ± 1241 (9) 635 ± 624 (8)

39,181 ± 60,025 (8) 12,982 ± 13,226 (13) 3575 ± 4478 (16) 1139 ± 2494 (29) 2304 ± 2651 (14) 2474 ± 3843 (8)

Class

Chloro phyceae

Diatomo phyceae

PAR

Temp. (°C)

NH4 + NH3 (lM)

pH Initial

Final NA NA 4.0 5.7 3.7 3.0 3.2 4.0 8.0 7.0 NA NA 7.0 NA 8.7 NA 9.2

Chla. reinhardtii Chlorella sp. Chlo. protothecoides Chlo. vulgaris Chlo. vulgaris Chlo. vulgaris Chlo. vulgaris Chlo. vulgaris A23 Chlo. vulgaris AA Chlo. vulgaris Dunaliella sp. D. salina D. tertiolecta CCAP D. tertiolecta UTEX N. pyriformis Sce. acuminatus Sce. obliquus

100 40–100 Dark 250 250 250 250 100 100 70 200–340 60 65 NA 35 200 10

25 23–29 28 18–22 18–22 18–22 18–22 28 ± 2 28 ± 2 20 ± 2 20 30 18 23 20 ± 1 NA 30

NA 7.2 6.1 5.9 5.2 4.5 3.8 7.0 7.0 7.0 NA 7.6 7.0 7.8 8.7 7.5 9.8

Anabaena azotica Arthrospira platensis M. aeruginosa 315 M. aeruginosa 905 Nostoc sp. Synechococcus sp. T. odorata

70 72 40 40 70 40 48–122

25 30 25 25 25 25 28

8.3 ± 0.2 8.0 8.3 ± 0.2 8.3 ± 0.2 8.3 ± 0.2 8.3 ± 0.2 7.0

Am. hyalina As. japonica Chaetoceros sp. Chaetoceros sp. Ch. affinis Cy. cryptica Cy. cryptica Gyrosigma spencerii Na. acceptata 6 Na. acceptata 8 Na. arenaria Na. cryptocephala Na. perminuta Na. salinarum Na. saprophila Nitzschia closterium Nitzschia dissipata Nitzschia dissipata Nitzschia dubiformis Nitzschia pungens Pl. aestuarii S. costatum S. costatum S. costatum S. sp. (Dokai Bay) S. sp. (Harima Nada)

100 200 NA 100 200 100 Dark 85 100 100 85 85 40 85 100 150 100 85 85 100 25 NA 200 100 100 100

22 20 NA 21 20 22 25 12 22 22 12 12 25 12 22 21 22 12 12 17 21 NA 20 17 21 21

7.5 NA 8.1–8.2 8.0 NA 7.5 7.5 8.0 (Tris) 7.5 7.5 8.0 (Tris) 8.0 (Tris) 8.0 8.0 7.5 8.1 7.5 8.0 8.0 NA NA 8.1 NA NA 8.0 8.0

10.4

7.0 NA NA NA NA NA NA 7.7 NA NA

NA 8.4 NA 9.0 NA NA NA NA NA 7.7 NA NA 9.4 NA

NH3 (lM)

n

EC50

EC50

NH4 + NH3 (lM)

NH3 (lM)

Control

100–10,000 5000–30,000 80,000–725000 2400–12,500 2400–12,500 2400–12,500 2400–12,500 10,000–71,428 10,000–143000 714–71,428 25–200 10–5000 250–16,000 1000–50,000 1–40 7000–70,000 900–4000

NA 60–360 80–725 1.2–6.3 1.2–6.3 1.2–6.3 1.2–6.3 70–500 70–1250 3–286 NA 0.3–1.5 1–48 40–2000 0.2–8 NA 730–3240

3 3 6 5 5 5 5 8 8 10 8 5 7 4 7 6 6

Stim Stim NS 8046 3805 6333 5622 31,622 56,234 NS NS Stim NS 4536 224 NA 3162

Stim Stim NS 3 3 3 3 210 562 NS NS Stim NS 225 2 NA 1258

1000–50,000 400–6400 500–10,000 500–10,000 500–10,000 500–50,000 350–35,000

100–5000 30–480 50–1000 50–1000 50–1000 50–5000 3–245

6 3 7 6 6 6 3

2448 Stim 3258 3735 1105 8961 Stim

388 Stim 316 251 466 1995 Stim

NO3

1000–3000 50–200 1–4000 50–1500 50–200 5–3000 1700–20,100 100–5000 1000–3000 5–3000 100–5000 100–5000 75–4000 100–5000 5–3000 286–1143 10–3000 100–5000 100–5000 220–880 1000–20,000 1–4000 50–200 220–880 10–1500 10–1500

14–42 NA NA 2–62 NA 0.1–42 31–362 2–100 14–42 0.1–42 2–100 2–100 4–216 2–100 0.1–42 12–49 0.1–42 2–100 2–100 NA NA NA NA NA 0.4–62 0.4–62

4 8 6 10 8 6 8 5 4 3 5 5 10 5 5 18 4 5 5 4 5 6 5 4 10 10

1224 NS 300 662 NS NS 243 1200 NS Stim 1256 2901 302 64 Stim 714 982 1609 1025 NS 4116 316 126 NS 423 135

14 NA NA 21 NA NS 126 25 NS NA 13 61 16 1 Stim NA 14 34 21 NA NA NA NA NA 14 5

NO3 NO3

NO3 NO3 NO3 NO3

NO3 NO3

NO3 NO3 NO3

NO3 NO3 NO3 NO3

NO3 NO3 NO3 NO3 NO3 NO3

NO3 NO3

NO3 NO3 NO3 NO3 NO3

References

Giordano et al. (2003) Kim et al. (2012) Shi et al. (2000) Pratt and Fong (1940) Pratt and Fong (1940) Pratt and Fong (1940) Pratt and Fong (1940) Przytocka-Jusiak et al. (1977) Przytocka-Jusiak et al. (1977) Tam and Wong (1996) Thomas et al. (1980) Norici et al. (2002) Fabregas et al. (1989) Chen et al. (2011) Källqvist and Svenson (2003) Park et al. (2010) Abeliovich and Azov (1976) Dai et al. (2008) Carvalho et al. (2004) Dai et al. (2008) Dai et al. (2008) Dai et al. (2008) Dai et al. (2008) Tan et al. (1993) Tadros and Johansen (1968) Thomas et al. (1980) Zgurovskaya and Kustenko (1968) Suksomjit et al. (2009a) Thomas et al. (1980) Tadros and Johansen (1968) Pahl et al. (2012) Admiraal (1977) Tadros and Johansen (1968) Tadros and Johansen (1968) Admiraal (1977) Admiraal (1977) Underwood and Provot (2000) Admiraal (1977) Tadros and Johansen 1968) Adams et al. (2008) Tadros and Johansen (1968) Admiraal (1977) Admiraal (1977) Bates et al. (1993) Rao and Sridharan (1980)) Zgurovskaya and Kustenko (1968) Thomas et al. (1980) Bates et al. (1993) Suksomjit et al. (2009a) Suksomjit et al. (2009a)


Cyano phyceae

Genus/species

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Table 2 Estimates of EC50 values (for growth, except where noted) for NH4 + NH3 (total ammonia) and for NH3 only calculated from final pH values for cultures growing at various PAR values, temperature and at a range of total ammonia concentrations. In some studies growth on NO3 was the control. Al: Alexandrium; Am: Amphiprora; As: Asterionella; Ch: Chaetoceros; Chla: Chlamydomonas; Chlo: Chlorella; Cyc: Cyclotella; Cyl: Cylindrotheca; D: Dunaliella; FW: freshwater; Gym: Gymnodinium; Gyr: Gyrodinium; H: Heterosigma; K: Karenia; L: Lingulodinium; M: Microcystis; N: Nephroselmis; Na: Navicula; NA: not available; NS: not significant; P: Prorocentrum; Pl: Pleurosigma; S: Skeletonema; Sce: Scenedesmus; Stim: stimulation of growth by ammonium; Syn: Synechocystis. T: Trentepolia; PAR: photosynthetically available radiation in lmol photons m2 s1.

NO3

Suksomjit et al. (2009a) Chang and Page (1995) Chang and Page (1995) Chang and Page (1995) Suksomjit et al. (2009a) NO3

19 NA NA NA 28 489 379 183 101 556 7 6 6 6 10 0.2–62 NA NA NA 0.4–62 5–500 27–875 27–875 27–875 10–1500 NA NA NA NA NA 21 15 15 15 21 Chattonella antiqua H. carterae H. carterae H. carterae H. akashiwo Raphido phyceae

100 160 80 40 100

8.0 NA NA NA 8.0

5–250 5–500 50–200 10–10,000 10–10,000 7–4000

20 20 20 NA 20 21 20 18 18 NA A. minutum A. minutum A. minutum Gym. splendens Gyr. instriatum K. mikimotoi L. polyedrum Peridinium sp. P. micans P. micans Dino phyceae

100 50 25 200–340 20 150 100 200 100 100 NA

NA NA NA NA NA 8.0 NA 6.0 6.0 8.1

NA NA NA 25–200 NA NA NA 7.2 7.2 7.7

6–200 6–200 6–200

NA NA NA NA NA 0.2–62 NA 0–3 0–3 NA

5 5 5 8 8 8 5 4 4 6

96 77 30 753(lag)NA NS 234 126 2264 2674 753

NA NA NA NO3 NA 9 NA 1 1 NA

NO3 NO3

Chang and McClean (1997) Chang and McClean (1997) Chang and McClean (1997) Thomas et al. (1980) Nagasoe et al. (2010) Suksomjit et al. (2009a) Thomas et al. (1980) Barker (1935) Barker (1935) Zgurovskaya and Kustenko (1968)


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and therefore the temperature effect has been taken into account in these calculations. For data in Table 2, the incubation duration ranged between 1 and 21 days. For those data sets where initial and final pH was documented, there were 8 experiments in which the pH decreased (6 in freshwater and 2 in seawater media) and 9 in which the pH increased (3 in freshwater and 6 in seawater). There was no pH change in 8 experiments, generally due to the use of Tris (Fabregas et al., 1989) or TAPS buffers (Dai et al., 2008), or very careful monitoring of pH (Tan et al., 1993; Tam and Wong, 1996; Källqvist and Svenson, 2003). The largest pH decrease was a 3 unit drop over 24 h during growth of Chlorella vulgaris under continuous light (Przytocka-Jusiak et al., 1977) and a 2.1 unit drop during growth of Chlorella protothecoides in darkness (Shi et al., 2000). The largest increase was 2.4 units during growth of Arthrospira platensis (Carvalho et al., 2004). In the later case, the final pH of 10.4, compounded by a temperature of 30 °C, led to an estimate of 90% ammonia in the medium. Otherwise, except for the studies of Abeliovich and Azov (1976), Adams et al. (2008), Källqvist and Svenson (2003) and Suksomjit et al. (2009a), the final pH values did not exceed 8.5, and remained in a range where the concentration of ammonia was very low and nearly negligible. A recent study (Dai et al., 2012) investigated the influence of ammonium on the Photosystem II (PSII) quantum yield of 18 freshwater algal species and ranked them according to their EC50 values in relation to the trophic state ranging from oligotrophic to hypertrophic. Overall, EC50 ranged from 260 to 500 lM for diatoms in socalled ‘‘oligo-’’ and ‘‘mesotrophic’’ regimes, from 560 to 1500 lM for cyanobacteria in eutrophic regimes, and from 1500 to 18,500 lM for chlorophytes in hypertrophic regimes. Quantum yield assesses short-term effects, but nevertheless, it reflects the longer term effects depicted in Table 2. Other variables, such as the Photosynthetically Available Radiation (PAR) could modulate the toxic effects of high ammonium levels and generally toxicity was higher at higher irradiances. Early work by Guillard (1963) indicated that ammonium inhibition of growth was higher in high light than in dim light. Similarly, Admiraal (1977) reported that inhibition of photosynthesis of estuarine benthic diatoms growing on 500 lM ammonium was enhanced at high PAR. This was confirmed by Hillebrand and Sommer (1996) who found 300 lM to be toxic to Pseudo-nitzschia multiseries at 230 lmol photons m2 s1, but not at 25 lmol photons m2 s1. On shorter time scales (minutes to hours), Drath et al. (2008) reported more inhibition of PS II activity in Synechocystis at 40 than at 10 lmol photons m2 s1. The sensitivity of Microcystis aeruginosa to ammonium increased with PAR in the range 50–1000 lmol photons m2 s1 (Dai et al., 2012). In contrast to these five studies above, two studies reported greater ammonium inhibition of growth at low PAR for A. minutum (Chang and Page, 1995) and Heterosigma akashiwo (Chang and McClean, 1997).

2.1. Acclimation and adaptation of growth rate to high ammonium in cultures Many of the laboratory experiments use species that have been in culture collections for many years and there is always concern about acclimation/adaptations that could have occurred in the culture collection. Berge et al. (2012) tested species that were maintained in culture collections for many years on enriched medium and it is likely that they experienced nutrient exhaustion and high pH frequently before they were transferred to new medium. They found that these long term strains tended to have lower growth rates and increased tolerance to high pH. Therefore, it is ideal to use recently isolated ecologically important species and simulated environmental factors to ensure applicability to field observations.

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Table 3 Induction of ammonium uptake by unicellular algae in laboratory cultures at various physiological N states and growing on various N sources. Lag = time to reach Vmax for ammonium uptake. PAR in lmol photons m2 s1. NA = not available. Nat: natural. Lim: limited; starv: starved; suff: sufficient. Species

PAR

T (°C)

pH

Previous N source

Physiological status

NH4 addition (lmol N/l)

Lag (h)

References

Chlamydomonas reinhardii Chlamydomonas reinhardii Chlamydomonas reinhardii Chlorella fusca Chlorella fusca Chlorella vulgaris Ditylum brightwellii Emiliana huxleyi Emiliana huxleyi Emiliana huxleyi Lingulodinium polyedrum Lingulodinium polyedrum Haslea ostrearia Nitzschia ovalis Nitzschia ovalis Phaeodactylum tricornutum Phaeodactylum tricornutum Platymonas striata Skeletonema costatum Skeletonema costatum Skeletonema costatum Thalassiosira antarctica Thalassiosira antarctica

65 65 500 70 140 Dark 175 150 150 150 175 100 100 100 nat PAR NA

25 25 25 25 25 25 15 13 13 13 15 18 NA NA

NO3 NH4 NH4 NH4NO3 NH4NO3 NH4NO3 NO3 NO3 NO3 NO3 NO3 NO3 NO3 NO3 NO3 NO3

N starv. 16 h N suff. N suff. N starv.16 h N starv. 16 h N starv. 16 h N.A. N suff. N suff. N suff. NA N starv. 24 h N suff. N suff. N starv. 24 h N suff.

900 200 200 9000 9000 10,000 1 5 10 17 3 10 30 30 40 200

0 0 0 48 24 0 75 85 60 72 24 0 48 48 24 96

Thacker and Syrett (1972) Cullimore and Sims (1980) Florencio and Vega (1983) Syrett and Morris (1963) ‘‘ ‘‘ ‘‘ Syrett and Fowden (1952) Eppley et al. (1969) Page et al. (1999) ‘‘ ‘‘ ‘‘ ‘‘ ‘‘ ‘‘ Eppley et al. (1969) Harrison (1976) Robert and Maestrini (1986) Robert and Maestrini (1986) Maestrini et al. (1986) ZoBell (1935)

200

20

7.0 7.0 7.0 7.4 7.4 6.0 NA NA NA NA NA NA 7.8 7.8 NA 7.6– 8.0 8.0

NH4

N suff.

300

0

Cresswell and Syrett (1982)

90 310 490 NA 50 50

20 17 17 13 0 0

NA NA NA NA NA NA

NH4 NA NH4 NO3 NO3 NO3

N N N N N N

1200 5 8 8 25 50

1 0 0 24 0 2

NA

Przytocka-Jusiak et al. (1977) acclimated an ammonium sensitive strain of Chlorella vulgaris originally maintained at 9860 lM, by culturing in medium containing inhibitory (53,570 lM), but not toxic, ammonium concentrations. The cells did not divide, but there was no lethal effect on cell survival. Four transfers at 7-d intervals were necessary to acclimate the cells to these excessively high ammonium concentrations. It was speculated that some cellular mechanism regulated the external pH so that it did not vary as much after acclimation (one unit or less over 24 h). In contrast, for the non-acclimated strain, the pH in the culture medium dropped by four units (8–4) over 24 h. Once Chlorella was acclimated, the C/N ratio exhibited values typical of N-limited cells at lower ammonium concentrations (12, 14.5 and 17.6 mol C mol1 N at 35,710, 17,860 and 9860 lM respectively and 6.3–7.9 at higher ammonium concentrations between 54,570 and 142,860 lM).

starv. 1 h suff. lim. starv. 450 lM. In contrast, higher ammonium may actually stimulate growth of other phytoplankters such as Synechococcus (Birdsey and Lynch, 1962; Kapp et al., 1975; Neilson and Larsson, 1980; Dai et al., 2008). Suksomjit et al. (2009b) compared the ammonium tolerance of Skeletonema isolated from the Seto Inland Sea (low ambient ammonium) with an isolate from Dokai Bay (ammonium 100 lM) and found that growth rate decreased at 500 lM in the former, but the Dokai Bay isolate tolerated ammonium up to 1500 lM. This suggests that Skeletonema from Dokai has adapted to the continually high ammonium concentrations of 100 lM in Dokai Bay. This is somewhat similar to the study of Han et al. (1992) who found that the volume specific photosynthetic rate of S. costatum in Tokyo Bay reached maximal values at around 140 lM ammonium and decreased at higher concentrations. Takahashi and Fukazawa (1982) using natural communities from Osaka Bay (Japan), reported that 50 lM enrichment stimulated the growth of Eutreptiella sp. (0.5 vs. 0.4 d1), Skeletonema costatum (1.0 vs. 0.7 d1) and Thalassiosira sp. (0.4 vs. 0.2 d1) relative to 10 lM, while the reverse was true for Gymnodinium sp. (0.15 vs. 0.45 d1) and Heterosigma sp. (0.5 vs. 0.3 d1). Domingues et al. (2011) reported specific net growth rates for several classes of phytoplankton in a tidal estuary following differential enrichments with ammonium (from 1 to 100 lM) and nitrate (100 lM) over several seasons. Overall, green algae growth was consistently stimulated by ammonium additions (up to 100 fold)). Cyanobacteria (reported in summer only) growth rates were strongly enhanced by high ammonium (up to 8-fold increase). Growth rates of diatoms decreased significantly in spring and summer with ammonium additions >50 lM, but this could be due to silicate limitation (estimated from data in Domingues et al., 2010). Dinoflagellates were clearly inhibited by 100 lM ammonium in the spring–summer transition and in summer (even reaching negative growth rates in summer). Ammonium appears to favor cyanobacterial dominance in lakes, possibly because of their superior uptake kinetics (Table 4), whereas nitrate enrichment may selectively stimulate growth of eukaryotes whose nitrate reductase can be more easily induced than that of cyanobacteria (Blomqvist et al., 1994). Similarly, Donald et al. (2011) found that ammonium and urea additions favored non-heterocystis cyanobacteria and chlorophytes at the expense of diazotrophic taxa. The total ammonia concentration from a sulphite pulp mill in northern Florida may reach >100 lM which was shown to be toxic to phytoplankton assemblages with reductions in algal biomass and species richness (Livingston et al., 2002). Microcosm experiments with the dominant diatom Skeletonema showed inhibition between 7 and 20 lM (i.e. surprisingly low compared to other studies) and major effects at >30 lM. For this system in Florida, it was recommended that total ammonia should not exceed 10 lM. High ammonium concentration was a selective factor in the distribution of benthic diatoms on an estuarine mudflat (Admiraal and Peletier, 1980). Navicula salinarum and Gyrosigma fasciola tolerated ammonium concentrations of 7000–10,000 lM, while four other species were less tolerant at 2000–7000 lM. 3. 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

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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 shortterm 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 not influence the outcome of species competition in the field that is expected to occur over several days of growth. In several studies, there appeared to be an induction or acclimation period for ammonium uptake at high ammonium concentrations (i.e. ammonium uptake does not reach Vmax immediately after the addition of ammonium). Table 3 summarizes the lag phase (or lack of it) that could be identified in several time series studies. Some lag phases were very long and up to 75–85 h in the early study by Syrett and Morris (1963) on Chlorella fusca (Morris, 1974). Possibly, the large amount of unbuffered ammonium chloride added (9000 lM) led to an immediate decrease in the ambient pH (Collos et al., 1992) that was deleterious to the initial uptake of ammonium or to growth in general. This time lag probably reflects an acclimation to high ammonium levels, with a very high interspecies variability, because overall, it was surprising that this time lag was not related to ammonium concentrations. For example, small ammonium additions (1–8 lM) still led to 24 h time lags for D. brightwellii (Eppley et al., 1969) and S. costatum (DeManche et al., 1979). Possibly, those long time lags were due to previous growth on nitrate. Conway (1977) showed that preconditioning on ammonium led to much smaller lags in ammonium uptake for S. costatum and in agreement with ZoBell (1935) and Cresswell and Syrett (1982) who worked on P. tricornutum. For species such as Chlorella that are very tolerant of high (10,000 lM) ammonium, there is still a significant acclimation period for the ammonium uptake system to reach its maximum capacity even when the cells were previously grown on ammonium (Syrett and Morris, 1963). This shows that there are detrimental effects of high ammonium levels on the ammonium uptake system. Page et al. (1999) mentioned the lack of ‘‘surge uptake’’ that is often observed and speculated that an ‘‘end-product regulator’’ might have shut down surge uptake. 3.1. Transition phases of uptake rates at high ammonium concentrations Table 4 summarizes the data on phase transitions (i.e. implementation of a second higher rate uptake system at higher substrate concentrations). This phenomenon of bi-phasic or multiphasic uptake systems is well known in higher plants (Clarkson and Lüttge, 1991), but less so in microalgae. The ratios of maximum uptake rates for high affinity systems (operating at low concentrations) to low affinity systems (at high concentrations) have been calculated from available published data. Such ratios in uptake are variable, but in some cases they can reach very high values, particularly among the Cyanophyceae, less so among the diatoms. Phytoplankton seem to exhibit such phase transitions across a wide variety of classes (cyanobacteria, diatoms, dinoflagellates) and habitats, ranging from prairie lakes to coastal lagoons (Table 4). In natural populations, these phase transitions led to a 2- to 6-fold increase in ammonium uptake over a rather narrow concentration range, and even more for Microcystis. This appears to be a nutrient acquisition strategy rather than a lack of cell control on uptake. Apart from P. delicatissima for which there are no data on growth rates, the highest ammonium concentrations in Table 4 have not been found to be toxic to growth of Alexandrium catenella (Collos et al., 2004), Chlorella vulgaris (Urhan, 1932), T. weissflogii (Conover, 1975), Microcystis (Dai et al., 2008) and Nannochloropsis (Hii et al., 2011). This enhanced uptake at high ammonium concentrations enables some species to acquire the limiting

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nutrient faster than other species and enables them to exploit transient elevated concentrations (Conway et al., 1976). This confers an ecological advantage over species that do not have a multi-phasic uptake system, as often observed in higher plants (Clarkson and Lüttge, 1991). 3.2. Cellular transport of ammonium and energetics The undissociated and uncharged ammonia molecule is lipid soluble and therefore it easily enters through the membrane and high intracellular concentrations can depolarize the membrane and possibly inhibit anion transport and affect the cell’s metabolism. In contrast, membranes are relatively impermeable to the charged ammonium ion. Internal pools of ammonium can reach up to 30% of the cell N content for A. catenella (Collos et al., 2006) and up to 66% of the cellular N content for T. weissflogii (Conover, 1975) without any apparent detrimental effect on cell growth. At typical cytoplasmic and vacuolar pHs of around 7–8 (Altenburger et al., 1991 and references therein; Taylor et al., 2012), ammonium would be the dominant form rather than ammonia. From an energetic point of view, ammonium should enhance growth relative to nitrate, especially at low light. However, contrary to expectations, Thompson et al. (1989) showed that higher growth rate only occurred at high light where energy is not limiting. As suggested by Collier et al. (2012), ‘‘it is possible that the predicted lower metabolic cost of ammonium compared to nitrate in terms of only N assimilation is offset by higher costs in terms of repair of photodamaged PSII and processes such as detoxification of oxygen radicals" (Bendixen et al., 2001). 3.3. Effects of ammonium inhibition Ammonium appears to be the ideal N source since its oxidation state eliminates the need for its reduction in the cell and thus it can be utilized immediately for the synthesis of amino acids. However, if the ammonium concentration is too high, it can be toxic and result in reduced growth. As observed in higher plants, the threshold of ammonium toxicity varies widely and thus there are sensitive and insensitive species. In the next section, some of the reasons for the occurrence of ammonium toxicity and its alleviation are explored for microalgae and some higher plants. 3.3.1. Direct Effects The most spectacular effect of high ammonium is cell lysis where algal cells burst immediately (Provasoli, 1958; Nagasoe et al., 2010), or within a few hours (Zgurovskaya and Kustenko, 1968) after ammonium addition. Early work on Chlorella vulgaris (Syrett, 1953; Hattori, 1957) indicated large increases in respiration upon addition of ammonium to nitrogen starved cells, and the oxygen absorbed was correlated with the ammonium taken up. Very elegant experiments using an isotope dilution method (Syrett, 1956) revealed that not only was more CO2 produced during ammonium assimilation than during nitrate assimilation, but CO2 was also fixed in darkness, in much greater proportions than that following nitrate addition. This was probably the first evidence for the anaplerotic pathway of carbon assimilation in microalgae. In the light, ammonium additions to ammonium limited or starved cultures often lead to a transient photosynthetic suppression for a few hours (Collos and Slawyk, 1979, 1984; Turpin, 1983; Elrifi and Turpin, 1985; Collos, 1986 and references therein). These suppressions are generally observed on time scales of 15 min (Turpin, 1983) to 6 h (Collos and Slawyk, 1979, 1984) and this is then followed by a stimulation of carbon fixation (Turpin, 1983;

Collos and Slawyk, 1984) relative to a control without ammonium addition. The time scale of ammonium perturbations is of utmost importance in the interpretation of such changes in carbon fixation results but remains totally unknown. Studies using high frequency (hourly) sampling revealed very large (50-fold, from 0.1 to 5 lM) and reproducible variations in ammonium concentrations between day and night over several diel cycles in the Newport River estuary (Litaker et al., 1988). The enclosures used by Parker et al. (2012b) led to the development of S. costatum with an initial ammonium level of about 10 lM, and this is consistent with the relatively high EC50 values for this species in Table 2. The apparent inhibition of ammonium uptake in the same study (p. 582, their Fig. 6) is similar to the ‘‘induction’’ phenomenon described in Table 3 when data are plotted as rates vs. concentrations, but the mechanism involved remains unexplained so far. Thus, it is unlikely that the 24 h incubations used by Parker et al., (2012a) could have led to such an apparent inhibition of primary production based on the results of the studies discussed above. Concerning the effect of ammonium on pigments, while the chlorophyll a (chla) content of Prorocentrum micans did not change between 1 and 71 lM NH4, that of S. costatum decreased by a factor of 2.5 between 1 and 71 lM and by a factor of 9 between 1 and 714 lM (Zgurovskaya and Kustenko, 1968), in parallel with decreases in photosynthesis (oxygen evolution). This was similar to the response of Nostoc sp. where a 10-fold decrease in chla content was reported between 1000 and 10,000 lM (Dai et al., 2008). In Dunaliella tertiolecta, the chla content first increased from 3 to 6 pg/cell between 250 and 1000 lM, then decreased from 6 to 3 pg/cell as ammonium further increased from 1000 to 32,000 lM, without effect on growth (Fabregas et al., 1989). This indicates that there are compensatory mechanisms to counteract detrimental effects on pigments. In Chlorella vulgaris, the chla content increased continuously from 0.1 to 2.1 pg/cell) between 714 and 71,428 lM (Tam and Wong, 1996), and also in Chlamydomonas reinhardtii (from 1.1 to 3.9 pg/cell) between 100 and 10,000 lM (Giordano et al., 2003). Thus, there are clear detrimental effects of high ammonium levels on the chla content of diatoms and cyanophyceae, that are also reflected in reduced growth rates (Table 2), but none in Chlorophyceae. In contrast, Leong and Taguchi (2004) reported an increase (4-fold) in chla /cell between 6 and 100 lM ammonium, and they attribute this to ‘‘low growth and larger cells due to the detrimental effect’’ of ammonium on growth. The free internal ammonium resulting from ammonium uptake is considered as source of stress (Giordano et al., 2003) by causing an intracellular pH disturbance (Britto and Kronzucker, 2002). Ammonium toxicity has long been thought to be due to uncoupling of photophosphorylation (Zhu et al., 2000; Britto and Kronzucker, 2006), but new evidence points to direct PS II photodamage due to ammonia binding to the Mn complex (Drath et al., 2008). 3.3.2. Indirect Effects The paradigm of pH changes brought about by ammonium uptake (Brewer and Goldman, 1976; Goldman and Brewer, 1980) is not really borne out by this review where a great variety of changes were observed (Table 2). The causes of pH changes in culture media during microalgal growth are probably complex and possibly also due to CO2/HCO3 uptake (Admiraal, 1977; Carvalho et al., 2004; Waser et al., 1998). Ammonium toxicity in higher plants has received extensive investigation and review relative to microalgae (Britto et al., 2001a,b; Britto and Kronzucker, 2002; Kronzucker et al., 2001). Ammonium in soils can range up to 2000–4000 lM and large scale forest decline has been linked to anthropogenic ammonium inputs and soil acidification (Pearson and Stewart, 1993; Britto and Kronzucker, 2002). Reasons for ammonium toxicity include, proton


extrusion, cytosolic pH disturbances displacement of crucial cations, and shifts in carbohydrate status. Sensitive plants such as barley show chlorosis of the leaves, significantly reduced growth, a decrease in cellular cations and a marked acidification in the hydroponic medium at 10,000 lM vs. 100 lM. Britto et al. (2001a,b) showed that at high ammonium concentrations, barley roots experience a breakdown in the regulation of ammonium influx, leading to excessive amounts of ammonium in the cytosol. This high ammonium concentration is pumped out of the cells at a high energetic cost (e.g. a 40% increase in root respiration) which may be responsible for the reduction in growth. This ammonium efflux may be up to 80% of the influx and has been viewed as futile transmembrane ammonium cycling, a new hypothesis to explain ammonium toxicity (Britto et al., 2001a,b; Kronzucker et al., 2001). In contrast, rice is insensitive to high ammonium because ammonium depolarizes the plasma membrane, whereas the potential difference in barley appears to be ammonium insensitive which leads to a high influx of ammonium. Ammonium toxicity has been shown to occur even in pH-buffered medium, suggesting that toxicity is not related only to changes in external pH and ammonium-induced cytosolic pH disturbance (Bligny et al., 1997). Recent work (Britto and Kronzucker, 2006) indicates that it is the energy-intensive nature of futile cycling of NH4 that leads to toxicity (high rate of ion cycling for low affinity transport systems). 3.4. Strategies to cope with high ammonium concentrations Cells have several strategies to tolerate high ammonium concentrations. In higher plants, it is not known why ammonium toxicity does not occur if both nitrate and ammonium are in the growth medium (Britto et al., 2001b). Ammonia and ammonium may be toxic to the cell since intracellular pools may be in the millimolar range (Britto et al., 2001a). Thus, ammonium toxicity may be alleviated by converting it quickly to amino acids (e.g. glutamine/glutamate) via glutamine synthetase that has a high affinity for ammonium (low Km) and glutamate dehydrogenase activities. One suggestion for the higher tolerance of green algae to high ammonium is that they have higher GS/GDH activities and hence the ammonium is converted quickly into amino acids, rather than accumulating in the cell (Klochenko et al., 2003). In the anaplerotic pathway, ammonium stimulates PEPCase that leads to rapid incorporation of ammonium into organic compounds to avoid toxicity (Giordano et al., 2003). In addition to the reduction in cytosolic ammonium by various synthesis pathways, ammonium may be transported to the vacuole where acid trapping of ammonium may occur. Ammonium may also be removed from the cytosol by ammonium extrusion/efflux to the medium (Britto et al., 2001a). In areas receiving high anthropogenic nitrogen loads, there is concern over the loss of seagrass beds and in some cases, the mass mortality of seagrasses has been suggested to be related to high ammonium (van Katwijk et al., 1997; Van der Heide et al., 2008). There are suggestions that ammonium may prevent an annual population from adopting a perennial reproduction strategy (van Katwijk et al., 1997). Ammonium concentrations may reach >200 lM with pHs of >9 in shallow estuaries due to anthropogenic inputs and algal biomass decomposition. Brun et al. (2002) assessed the toxicity of ammonium pulses on the survival and growth of Zostera noltii and found that the toxic effect depended on the internal carbon balance between photosynthesis and sucrose reserves that are needed to sustain nitrogen assimilation and the conversion into amino acids. Normally, there is an above and below ground mobilization of sucrose to meet increased carbon demands arising from ammonium assimilation. Brun et al. (2002) found that a repeated large pulse of ammonium quickly in-

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creased nitrogen transport and assimilation, and carbon demands and drained the carbon pool, resulting in intracellular accumulation of ammonium and hence toxic effects to photosynthesis and growth. At pH 8, Zostera marina showed no effects of high ammonium concentrations up to 150 lM, but at pH 9, it became necrotic in a few days (van der Heide et al., 2008). Remarkably, with high shoot densities at pH 9, there was no necrosis and toxic effects appeared to be alleviated through ‘joint’ ammonium uptake occurring under high shoot densities. Thus, ammonium toxicity was dependent on shoot density. In another study with Z. marina, van Katwijk et al. (1997) observed ammonium toxicity at 125 lM and even down to 25 lM. They found that low light enhanced ammonium toxicity because photosynthesis was reduced along with carbon stores that are needed for the assimilation of ammonium into amino acids. Similarly, Villazán et al. (2013) found that elevated ammonium and low light formed a deadly combination for Z. marina that explained why eelgrass and other seagrasses deteriorate under nitrogen-rich, low light conditions. Higher temperatures (15 vs. 20 °C) increased toxicity due to higher uptake rates of ammonium and higher carbon-consuming respiration rates. Eelgrass leaves are more susceptible to toxicity than roots, possibly because leaves have a higher uptake rate and the surrounding seawater pH is 8.2, compared to pH of 7.5 in sediments, despite the much higher ammonium concentrations in sediments of between 100–1000 lM. In a comparative study of two duckweeds, Monselise and Kost (1993) found that Lemna gibba from a high ammonia site was more efficient at detoxifying excessive ammonia by simply accelerating the usual GS-GOGAT nitrogen assimilation system (i.e. an ‘‘ammonium trapping’’ process), producing glutamine and c-aminobutyrate and leaving little intracellular free ammonia/ammonium. In contrast, Wolffia arrhiza from a low ammonium site had a less efficient ammonium assimilation system, probably due to a relative shortage of glutamate. This suggestion was supported by the addition of a-ketoglutarate which resulted in a significant decrease in intracellular ammonia. Thus, a-ketoglutarate is a precursor for the ammonium trapping mechanism. Substantial evidence indicates that d-amides are key detoxification products when plants are exposed to high levels of total ammonia and they act as storage reservoirs or sinks for intracellular ammonia/ammonium which may be toxic for cells.

4. 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, 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 influence ammonium and metal toxicity and other components of the effluent (Hogan et al., 2005). In Australia, effluent discharge from Melbourne’s sewage treatment plant caused a decrease in brown macroalgae and an increase in opportunistic green macroalgae near the discharge site. Subsequent chronic toxicity testing using the microalgal growth rate of Nitzschia closterium, macroalgal germination and cell division of Hormosira banksii, larval development of the scallop Chlamys asperrima, confirmed that ammonium was the major cause of effluent toxicity (Hogan et al., 2005; Adams et al., 2008). Macroalgal cell division and germination was the most sensitive test (EC50 = 100 lM), followed by scallop larval development (EC50 = 200 lM) and finally the microalgal growth inhibition test (EC50 = 700 lM). Comparisons of toxicity using calculated unionized ammonia (NH3) and

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total ammonia gave similar results. High ammonium toxicity values for four marine amphipods ranged from a EC50 of 3000 to 10,000 lM, probably because amphipods are exposed to higher ammonium concentrations associated with the sediments as compared to the water column (Kohn et al., 1994). In a very carefully designed bioassay, Källqvist and Svenson (2003) used the chlorophyte Nephroselmis pyriformis since it was the most sensitive of the nine species that they tested. They used 3 mM HEPES buffer to control pH and removed NH3 by raising the pH to >11 with vigorous bubbling for >2 h. This is one of the very few studies that has evaluated the toxicity of ammonium and ammonia separately. Most other studies have assessed the combined effect of ammonia and ammonium (i.e. total ammonia) and since toxicity increased with increasing pH, these studies inferred that the toxicity was mainly due to ammonia (e.g. Hogan et al., 2005; Adams et al., 2008). Källqvist and Svenson (2003) found that ammonia had an EC50 of 2.4 lM, while the EC50 for ammonium was 100 times higher at 224 lM and the EC50 for total ammonia at pH = 8.0 was 71 lM. They concluded that there is a joint toxicity effect (i.e. both ammonia and ammonium), but that ammonium is much less toxic, possibly because transport of this charged ion through the membrane is restricted compared to the passive entry of the uncharged ammonia molecule. There are no guidelines for ammonium in sediments. The Australian and New Zealand guidelines (termed ‘trigger values’) for marine and estuarine waters are as follows : 65 lM total ammonia for 95% species protection (at 20 °C and pH 8), derived from chronic no observable effects (ANZECC/ARMCANZ, 2000). Canada has no recommended guideline for ammonium in marine waters (CCME, 2000). The EU directive for freshwater has guidance and imperative values of 14.3 and 71.5 lM, respectively. The USEPA has a chronic marine criterion of 54.3 lM total ammonia (USEPA, 1989). Batley and Simpson (2009) revised the trigger value (i.e. no further action is required) for ammonia using a greatly expanded toxicity data set and their new 95% species protection value was 32.9 lM, about half the previous value of 65 lM. For sediment pore waters, a trigger value of 300 lM was recommended. They stated that a guideline concentration of 565 lM (i.e. the concentration below the EC50 for 95% of the species), represents a major risk of acute toxicity to sensitive species. 5. Other examples of NH4 toxicity

5.1. 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. Nannochloropsis sp. appears to be an ideal species since it can tolerate up to 900 lM ammonium and grows significantly faster on ammonium than nitrate (Hii et al., 2011). In fish ponds, there is a daily pH cycle due to algal photosynthesis and the removal of CO2. The pH may increase to >9 in the late afternoon, causing a >50-fold increase in ammonia (Wurtz, 2003). Since the %NH3 is pH dependent, the fish must endure toxic ammonium levels for a few hours and the pond may heat up in the late afternoon, causing a further small increase in the ammonia concentration. Ammonia may be produced in/near the sediments due to the decomposition of the algae and the uneaten fish food and depending on the aeration system, it

may diffuse from the sediments or be mixed up into the water column. Another ammonia sink besides algal uptake is nitrification, the bacterial oxidation of NH3 to NO 2 and NO3 . Ammonia concentrations tend to be higher in the winter (180 to 285 lM) in the fish ponds and lowest in summer (35 lM) due to the highest algal growth in summer (Hargreaves and Tucker, 2004). The complex nitrogen biogeochemistry of aquaculture ponds has been well reviewed by Hargreaves (1998). A common problem in many ponds is the occurrence of the toxin-producer Prymnesium parvum. For some fish, ammonium sulphate can be added to an ammonia concentration of 10–20 lM that will kill P. parvum, but minimize mortality of several species of bass (Barkoh et al., 2004). Of course, the safe level for fish varies greatly with species, size and life stage. The average of the mean acute toxicity values for 32 freshwater fish is 200 lM compared to 130 lM for marine fish, indicating that marine species are slightly more sensitive to ammonia toxicity that freshwater species (Randall and Tsui, 2002). 5.2. Removal of ammonia in wastewater treatment ponds Ammonia is one of the major constituents of domestic wastewater and concentrations range from 10–200 mg L1 (1000– 20,000 lM) (Konig et al., 1987; Thomas et al., 1980). There has been considerable interest in using high rate oxidation ponds to treat wastewater (Abeliovich and Azov, 1976). One alga that is particularly tolerant to ammonium is Chlorella vulgaris and hence it has a potential role in the removal of ammonium from wastewater effluent since it grows well at concentrations ranging up to 20,000 lM (Przytocka-Jusiak et al., 1984; Tam and Wong, 1996; Kim et al., 2010). The freshwater alga, Scenedesmus obliquus, has a high growth rate in laboratory cultures (5–8 h doubling time), but in the sewage ponds, a retention time of >100 h was necessary to maintain a low ammonium concentration because growth was slow when ammonium concentrations were >2000 lM and pH was >8 (Abeliovich and Azov, 1976). In follow-up experiments, Azov and Goldman (1982) observed that when the pH rises to 9.5 at 25 °C, 2000–3000 lM total ammonia led to a 50–90% reduction in photosynthesis. At the higher temperature of 25 °C, only one-third the total ammonia is required to produce the same free ammonia as at 10 °C and 15 times less total ammonia is required at pH 10 as at pH 8. Therefore, it is essential to know the pH and temperature of the pond from which the concentration of ammonia can be calculated and not just the total ammonia concentration. When algal systems are used for wastewater treatment and the source of N cannot be regulated, pH control is necessary to avoid algal ammonia toxicity (Azov and Goldman, 1982). Goldman et al. (1982c) compared some freshwater and marine species and concluded that marine species cannot tolerate pH > 9.5, except for Phaeodactylum tricornutum which grew at a pH of 10.3. The unique ability of P. tricornutum to grow at high pHs is a major factor that explains its frequent dominance in large-scale outdoor cultures (Goldman et al., 1982c). Therefore, P. tricornutum would appear to be an ideal/unique species for further studies on the interaction of pH and ammonium toxicity. Of the freshwater species, Scenedesmus obliquus grew better at a pH of 10.6 than Chlorella vulgaris. Duckweed systems are another option for sustainable wastewater treatment because they have high growth rates and high nutrient content that can be used for animal feeds (Körner et al., 2001). At low total ammonia concentrations of 9 where ammonia is more abundant, and these high pHs are seldom observed in the field. The ranking of algal groups observed in the cultures is somewhat reproduced in the field, although such comparisons are fraught with difficulties such as long-term acclimation of strains in cultures (Berge et al., 2012) and problems with identification of species such as ‘‘Chlorella-like’’ cells in the lower size classes in field samples. There is a need for more well designed laboratory experiments, similar to the excellent study by Källqvist and Svenson (2003) to assess the toxicity of ammonium separately from ammonia over a range of pHs, temperatures, irradiances and over a range of ecologically relevant ammonium concentrations. More studies are needed for recently isolated ecologically important species and an assessment of environmental factors, especially pH and any multiplicative effects with temperature, light, and salinity. 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. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.marpolbul.201 4.01.006. References Abeliovich, A., Azov, Y., 1976. Toxicity of ammonia to algae in sewage oxidation ponds. Appl. Environ. Microbiol. 31, 801–806.

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Further reading section Ahmad, I., Hellebust, J.A., 1984. Nitrogen metabolism of the marine microalga Chlorella autotrophica. Plant Physiol. 76, 658–663. Anderson, D.M., Kulis, D.M., Binder, B.J., 1984. Sexuality and cyst formation in the dinoflagellate Gonyaulax tamarensis: cyst yield in batch cultures. J. Phycol. 20, 418–425. Baek, S.H., Shimode, Han, M., Kikuchi, T., 2008. Growth of dinoflagellates, Ceratium furca and Ceratium fusus in Sagami Bay, Japan: the role of nutrients. Harmful Algae 7, 729–739. Barkoh, A., Smith, D.G., Schlechte, J.W., 2003. An effective minimum concentration of un-ionized ammonia nitrogen for controling Prymnesium parvum. North Am. J. Aquacult. 65, 220–225. Boussiba, S., 1989. Ammonia uptake in the alkalophilic cyanobacterium Spirulina platensis. Plant Cell Physiol. 30, 303–308. Bruno, S.F., McLaughlin, J.J.A., 1977. The nutrition of the freshwater dinoflagellate Ceratium hirundinella. J. Protozool. 24, 548–553. Converti, A., Scapazzoni, S., Lodi, A., Carvalho, J.C.M., 2006. Ammonium and urea removal by Spirulina platensis. J. Ind. Microbiol. Biotechnol. 33, 8–16. de-Bashan, L.E., Trejo, A., Huss, V.A.R., Hernandez, J.-P., Bashan, Y., 2008. Chlorella sorokiniana UTEX 2805, a heat and intense, sunlight-tolerant microalga with potential for removing ammonium from wastewater. Bioresour. Technol. 99, 4980–4989. Dixon, G.K., Syrett, P.J., 1988. The growth of dinoflagellates in laboratory cultures. New Phytol. 109, 297–302. Eker-Develi, E., Kideys, A.E., Tugrul, S., 2006. Effect of nutrients on culture dynamics of marine Phytoplankton. Aquat. Sci. 68, 28–39. Fidalgo, J.P., Cid, A., Abalde, J., Herrero, C., 1995. Culture of the marine diatom Phaeodactylum tricornutum with different nitrogen sources: growth, nutrient conversion and biochemical composition. Cah. Biol. Mar. 36, 165–173. Fukao, T., Nishijima, T., Yamaguchi, H., Adachi, M., 2007. Utilization of urea by the 6 species of red tide phytoplankton. Bull. Plankton Soc. Jpn. 54, 1–8. Grant, B.R., Madgwick, J., DalPont, G., 1967. Growth of Cylindrotheca closterium var. californica (Mereschk) Rieman & Lewin on nitrate, ammonia, and urea. Aust. J. Mar. Freshw. Res. 18, 129–135. Hamasaki, K., Horie, M., Tokimitsu, S., Toda, T., Taguchi, S., 2001. Variability in toxicity of the dinoflagellate Alexandrium tamarense isolated from Hiroshima Bay, western Japan, as a reflection of changing environmental conditions. J. Plankton Res. 23, 271–278. Herndon, J., Cochlan, W.P., 2007. Nitrogen utilization by the raphidophyte Heterosigma akashiwo: growth and uptake kinetics in laboratory cultures. Harmful Algae 6, 260–270. Hosaka, M., 1992. Growth characteristics of a strain of Heterosigma akashiwo (Hada) isolated from Tokyo Bay, Japan. Bull. Plankton Soc. Jpn. 39, 49–58. Iwasaki, H., Sasada, K., 1969. Studies on the red tide dinoflagellates. II. On Heterosigma inlandia appeared in Gokasho Bay, Shima Peninsula. Bull. Jpn. Soc. Sci. Fish. 35, 943–947.

Y. Collos, P.J. Harrison / Marine Pollution Bulletin 80 (2014) 8–23 Kratz, W.A., Myers, J., 1955. Nutrition and growth of several blue-green algae. Am. J. Bot. 42, 282–287. Lee, Y.S., 2008. Utilization of various nitrogen, phosphorus, and selenium compounds by Cochlodinium polykrikoides. J. Environ. Biol. 29, 799–804. Leong, S.C.Y., Murata, A., Nagashima, Y., Taguchi, S., 2004. Variability in toxicity of the dinoflagellate Alexandrium tamarense in response to different nitrogen sources and concentrations. Toxicon 43, 407–415. Lim, P.T., Leaw, C.P., Kobiyama, A., Ogata, T., 2010. Growth and toxin production of tropical Alexandrium minutum Halim (Dinophyceae) under various nitrogen to phosphorus ratios. J. Appl. Phycol. 22, 203–210. Lindström, K., 1991. Nutrient requirements of the dinoflagellate Peridinium gatunense. J. Phycol. 27, 207–219. Liu, M.S., Hellebust, J.A., 1974. Uptake of amino acids by the marine centric diatom Cyclotella cryptica. Can. J. Microbiol. 20, 1109–1118. McLaughlin, J.J.A., 1958. Euryhaline chrysomonads: nutrition and toxigenesis in Prymnesium parvum, with notes on Isochrysis galbana and Monochrysis lutheri. J. Protozool. 5, 75–81. Nakamura, Y., Watanabe, M.M., 1983. Growth characteristics of Chattonella antiqua. Part 2. Effects of nutrients on growth. J. Oceanogr. Soc. Jpn. 39, 151–155. Norris, L., Chew, K.K., 1975. Effect of environmental factors on growth of Gonyaulax catenella. In: LoCicero, V.R. (Ed.), Proceedings of the First International Conference on Toxic Dinoflagellate Blooms, November, 1974, Boston, Massachusetts. Massachusetts Science and Technology Foundation, Massachusetts, pp. 143–152. Paone, D.A.M., Stevens Jr., S.E., 1981. Nitrogen starvation and the regulation of glutamine synthetase in Agmenellum quadruplicatum. Plant Physiol. 67, 1097– 1100. Perez-Garcia, O., Bashan, Y., Puente, M.E., 2011. Organic carbon supplementation of sterilized municipal wastewater is essential for heterotrophic growth and removing ammonium by the microalga Chlorella vulgaris. J. Phycol. 47, 190–199. Pintner, I.J., Provasoli, L., 1958. Artificial cultivation of a red-pigmented marine blue-green alga, Phormidium persicinum. J. Gen. Microbiol. 18, 190–197. Provasoli, L., McLaughlin, J.J.A., 1963. Limited heterotrophy of some photosynthetic dinoflagellates. In: Oppenheimer, C.H. (Ed.), Symposium in Marine Microbiology. Thomas C.C., Springfield, Ill., pp. 105–113. Rao, V.N.R., Ragothaman, G., 1978. Studies on Amphora coffeaeformis II. Inorganic and organic nitrogen and phosphorus sources for growth. Acta Bot. Ind. 6, 146– 154.

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Samejima, H., Myers, J., 1958. On the heterotrophic growth of Chlorella pyrenoidosa. J. Gen. Microbiol. 18, 107–117. Shilo, M., Shilo, M., 1962. The mechanism of lysis of Prymnesium parvum by weak electrolytes. J. Gen. Microbiol. 29, 645–658. Singh, H.N., Srivastava, B.S., 1968. Studies on morphogenesis in a blue-green alga. I. Effect of inorganic nitrogen sources on developmental morphology of Anabaena doliolum. Can. J. Microbiol. 14, 1341–1346. Siu, G.K.Y., Young, M.L.C., Chan, D.K.O., 1997. Environmental and nutritional factors which regulate population dynamics and toxin production in the dinoflagellate Alexandrium catenella. Hydrobiology 352, 117–140. Soletto, D., Binaghi, L., Lodi, A., Carvalho, J.C.M., Converti, A., 2005. Batch and fedbatch cultivations of Spirulina platensis using ammonium sulphate and urea as nitrogen sources. Aquaculture 243, 217–224. Spoehr, H.A., Milner, H.W., 1949. The chemical composition of Chlorella; effect of environmental conditions. Plant Physiol. 24, 120–149. Su, H.M., Chiang, Y.-M., Liao, I.-C., 1993. Role of temperature, salinity and ammonia on the occurrence of the Taiwanese strain of Alexandrium tamarense. In: Smayda, T.J., Shimizu, Y. (Eds.), Toxic Phytoplankton Blooms in the Sea. Elsevier, pp. 837–842. Tadros, M.G., Johansen, J.R., 1988. Physiological characterization of six lipidproducing diatoms from the southeastern United States. J. Phycol. 24, 445–452. Tai, L.-S., 1934. On the cultivation of a photosynthetic dinoflagellate (Ceratium sp.). Chin. J. Physiol. 8, 111–118. Watanabe, M.M., Nakamura, Y., Mori, S., Yamochi, S., 1982. Effects of physicochemical factors and nutrients on the growth of Heterosigma akashiwo Hada from Osaka Bay, Japan. Jpn. J. Phycol. 30, 279–288. Wehr, J.D., Brown, L.M., O’Grady, K., 1987. Highly specialized nitrogen metabolism in a freshwater phytoplankter, Chrysochromulina breviturrita. Can. J. Fish. Aquat. Sci. 44, 736–742. Wood, G.J., Flynn, K.J., 1995. Growth of Heterosigma carterae (Raphidophyceae) on nitrate and ammonium at three photon flux densities: evidence for N stress in nitrate-growing cells. J. Phycol. 31, 859–867. Yamaguchi, M., Itakura, S., Uchida, T., 2001. Nutrition and growth kinetics in nitrogen- or phosphorus-limited cultures of the ‘novel red tide’ dinoflagellate Heterocapsa circularisquama (Dinophyceae). Phycologia 40, 313–318.

Nitrobacter winogradskyi transcriptomic response to low and high ammonium concentrations.

Cadmium toxicity to freshwater algae.

[Concerning the question of anaerobic growth in unicellular green algae].

Amines in unicellular green algae. 2. Amines in Scenedesmus acutus.

Anilines: selective toxicity to blue-green algae.

Improved methane production from brown algae under high salinity by fed-batch acclimation.

Sublethal toxicity of chlorpyrifos to salmonid olfaction after hypersaline acclimation.

Toxicity of ammonia to algae in sewage oxidation ponds.

Acclimation to high temperature during pollen development.

Acute toxicity of high concentrations of carbon dioxide in rats.

Preparation of photosynthetically active particles from synchronized cultures of unicellular algae.

Effect of relative humidity on the survival of airborne unicellular algae.

Cryptomonad algae are evolutionary chimaeras of two phylogenetically distinct unicellular eukaryotes.

Effects of acclimation to altitude on oxygen affinity and organic phosphate concentrations in pigeon blood.

The role of exopolymeric substances in the bioaccumulation and toxicity of Ag nanoparticles to algae.

Three dimensional quantitative structure-toxicity relationship modeling and prediction of acute toxicity for organic contaminants to algae.

Toxicity of ciprofloxacin and sulfamethoxazole to marine periphytic algae and bacteria.

Pollen Development at High Temperature: From Acclimation to Collapse.

Toxicity of fuel-contaminated soil to Antarctic moss and terrestrial algae.

Comparative toxicity of 20 herbicides to 5 periphytic algae and the relationship with mode of action.

Evaluation of toxicity data to green algae and relationship with hydrophobicity.

Size-dependent toxicity of silver nanoparticles to bacteria, yeast, algae, crustaceans and mammalian cells in vitro.

Plant biomechanics: High-endurance algae.

Toxicity evaluation of ammonium sulphate and urea to three developmental stages of freshwater snails.