Cita [6] Aclimatación y toxicidad de las altas concentraciones de amonio.

8/12/2019 Cita [6] Aclimatacin y toxicidad de las altas concentraciones de amonio.

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Review

Acclimation and toxicity of high ammonium concentrations

to unicellular algae

Yves Collos a,, Paul J. Harrison b

a Ecologie des Systmes Marins Ctiers (UMR5119), Universit Montpellier 2, CNRS, IRD, case 093, 34095 Montpellier Cedex 5, Franceb 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

EC50for 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 mic-

roalgae suggests the following rankings. Mean optimal ammonium concentrations were 7600, 2500,

1400, 340, 260, 100 lM for Chlorophyceae, Cyanophyceae, Prymnesiophyceae, Diatomophyceae, Raphid-ophyceae, 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 Cyanophy-

ceae, Dinophyceae, Diatomophyceae, and Raphidophyceae. Ammonia toxicity is mainly attributed to NH3at pHs >9 and at pHs


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Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

Appendix A. Supplementary material. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

1. 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 environ-

ments, 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 lM ammonium was toxic for 200 spe-

cies or clones of oceanic phytoplankton as determined by fluores-cence. 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,b, 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 re-

ported 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 dif-

ferent 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 ammo-

nium 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.

1.1. Influence of pH, temperature and salinity on ammonia/ammonium

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 mem-

branes. Since it is a gas, it is volatile and can be lost to the atmo-

sphere, 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 meth-

ods 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 salin-ity has only a minor influence (Whitfield, 1974; Emerson et al.,

1975; Bower and Bidwell, 1978; Spotte and Adams, 1983). In gen-

eral, at the pH of seawater at 8.0 and 20o

C, only about 10% of thetotal 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 con-

centration 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 aquacul-

ture 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 %NH3for 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 risein temperature over the 030 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 hard-

ness 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


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causes an increase in the pH due to the release of OH- ions and dur-

ing growth on urea, there is little change in the pH ( Raven and

Smith, 1976; Goldman et al., 1982a,b,c; Raven, 1988; Britto et al.,

2001b).

1.2. Nitrogen cycle: ammonium sources and sinks

Generally, ammonium concentrations in surface waters rangeupto3 lM. In contrast, anthropogenic inputs of ammonium fromatmospheric deposition, agricultural activities and sewage are re-

garded as new nitrogen (Dugdale and Goering, 1967). Examples

of anthropogenically produced atmospheric sources are transpor-

tation emissions, and volatilization from manure produced in ani-

mal 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 1040% 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 in-

crease 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 re-

garded as regenerated N and is produced by excretion from ani-

mals and by bacterial regeneration/recycling of organic N

compounds in the sediments and water column (Galloway et al.,

2003, 2004). Another minor source of ammonium in sediments

that is produced via nitrogen cycle processes is dissimilatory ni-

trate reduction (DNR). In sediments with high sulphide concentra-

tions, 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 reac-tions 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 phyto-

plankton 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).

1.3. High ammonium levels in various environments

The highest values of ammonium are generally found in fresh-

water environments. Maxima ranged from 100 lM in Wascana

Lake, Canada (Donald et al., 2011), 200 lM in prairie lakes, Canada(Murphy and Brownlee, 1981), 300 lM in Lake Taihu, China (Chenet al., 2003) and up to 680 lM recorded in summer 1990 in LakeLittle Mere, UK (Carvalho, 1994). Maxima sometimes increased

over time such as in Lake Donghu, China from 10lM in the1950s to 60lM in the 1980s and up to 100 lM in 2001 (Daiet al., 2012). In lagoons, maxima ranged from 9 lM in Thau lagoon,France (Collos et al., 2005), to 300 lM in Bolmon lagoon, France(Chomrat et al., 2007) and Balearic islands lagoons (Lucena-Moya

et al., 2012) and up to 400 lM in Portuguese lagoons (Coutinhoet al., 2012).

In estuaries, ammonium concentrations are generally related to

salinity (S), with high values in the less saline part, such as 300lMin the Ems-Dollard estuary (Admiraal, 1977), >400 lM inDeep Bay,

Hong Kong (Xu et al., 2010) or 1000 lM in the Colne estuary (UK)at S= 10, decreasing to 100lM at S =20 and 20 lM at S= 30

(Underwood and Provot, 2000). In coastal waters, the highest val-

ues are found near sewage outfalls: >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 (Thomasand Carsola, 1980), 3000lM 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 Frhi, 1999).While temporal increases can be found in some marine environ-

ments such as in Osaka Bay, increasing from 30lM in 1980 (Yamochiand Abe, 1984) to 300 lM in 2007 (Yamamoto et al., 2010), othermaxima seem to be more stable such as in Suisun Bay (USA):

27 lM in 1974 (Glibert, 2010), 16 lM in 20002003 (Wilkersonet al., 2006), 14 lM in 2006 (Parker et al., 2012a,b). High frequencysampling has revealed large diel changes in ammonium concentra-

tions that could increase 50-fold (from 0.1 to 5 lM) during thenight 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. 2shows 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 cer-

tain concentration. A similar experiment was conducted with ten

species of estuarine benthic diatoms and good growth occurred

at 17,000 lM nitrate, 100010,000lM nitrite, but only 500lMammonium was inhibitory (Rao and Sridharan, 1980). It is interest-

ing 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 highergrowth 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 mar-

ine 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 aqua-

tic 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 fromRao and Sridharan, 1980). Ammo-nium (triangles), nitrite (squares), nitrate (diamonds).

10 Y. Collos, P.J. Harrison/ Marine Pollution Bulletin 80 (2014) 823


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the medium and its potential toxicity. Some studies used pH buf-

fers to keep pH relatively constant during batch culture growth.

In addition, many studies did not use ecologically important spe-

cies. Data were also examined for indications of possible acclima-

tion of unicellular algae to high ammonium levels and the%NH3at known pHs.

Names of microalgae were checked against the Algaebase

(http://algaebase.org/) or the WoRMS (World Register of MarineSpecies) database and currently accepted names were used when-

ever 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 maxi-

mal 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 concen-

tration 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 notedthis 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 concentra-

tion is the concentration at which no growth was observed.

2. Effect of ammonium on growth rates of unicellular algae

Table 1summarizes optimum, inhibitory and toxic ammonium

concentrations (details inSupplementary tables) for growth rates

of six classes of unicellular algae. The ranking of the six algal clas-

ses in terms of their tolerance to high ammonium levels was as

follows:Chlorophyceae > Cyanophyceae > Diatomo

phyceae > Raphidophyceae > Prymnesiophyceae > Dinophyceae.

The KruskalWallis test and Dunns multiple comparison test

revealed that for toxic concentrations, Chlorophytes were signifi-

cantly 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 dino-

flagellates (p< 0.01). In general, comparing the six classes, dino-

flagellates 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

Ochromonassp. at 100 lM ammonium and Watson and McCauley(2005) found optimal concentrations at 140lM for Uroglenopsisamericana and Synura petersenii and at 14 lM for Dinobryoncylindricum. 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; Loureno

et al., 2002), with toxic ammonium levels of 5000, 1200 and

11,000 lM for Hemiselmis virescens, Hillea sp., and Rhodomomassp. respectively.

Concerning the Dictyochophyceae, optimal growth ofAureococ-

cus anophagefferensoccurred at 50lM ammonium, with inhibitoryand 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 ammo-

nium concentrations and their effects on growth rates (Table 2).

EC50values 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 fromthe one derived from Table 1: Chlorophyceae > Cyanophy-

ceae > Diatomophyceae > Dinophyceae > Raphidophyceae, but the

dinoflagellates, the diatoms and the raphidophytes were not signif-

icantly 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 EC50values for ammonia toxicity for the five algal groups. We found

the following ranking: Cyanophyceae > Chlorophyceae > Diatomo-

phyceae > 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 dinoflagel-

lates (p< 0.01).

Many of the effects of ammonia on growth rates were deter-

mined 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). There-

fore, growth rates at lower temperatures should be less sensitive

to a certain total ammonium concentration than at higher temper-

atures since NH3is more toxic than NH4 . 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 Mean SD Mean SD

Optimal Inhibitory Toxic

Chlorophyceae 7572 7619 (17) 23,758 25,922 (14) 39,181 60,025 (8)

Cyanophyceae 2486 1570 (10) 6616 6514 (10) 12,982 13,226 (13)

Diatomophyceae 337 409 (27) 725 839 (27) 3575 4478 (16)

Dinophyceae 110 77 (23) 324 283 (25) 1139 2494 (29)

Prymnesiophyceae 1432 1197 (12) 958 1241 (9) 2304 2651 (14)

Raphidophyceae 263 332 (8) 635 624 (8) 2474 3843 (8)

Y. Collos, P.J. Harrison/ Marine Pollution Bulletin 80 (2014) 823 11
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and therefore the temperature effect has been taken into account

in these calculations.

For data inTable 2, the incubation duration ranged between 1

and 21 days. For those data sets where initial and final pH was doc-

umented, there were 8 experiments in which the pH decreased (6

in freshwater and 2 in seawater media) and 9 in which the pH in-

creased (3 in freshwater and 6 in seawater). There was no pH

change in 8 experiments, generally due to the use of Tris (Fabregaset al., 1989) or TAPS buffers (Dai et al., 2008), or very careful mon-

itoring of pH (Tan et al., 1993; Tam and Wong, 1996; Kllqvist and

Svenson, 2003). The largest pH decrease was a 3 unit drop over

24 h during growth ofChlorella vulgaris under continuous light

(Przytocka-Jusiak et al., 1977) and a 2.1 unit drop during growth

ofChlorella 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), Kllqvist and

Svenson (2003)andSuksomjit et al. (2009a), the final pH values

did not exceed 8.5, and remained in a range where the concentra-

tion 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 fresh-

water algal species and ranked them according to their EC50values

in relation to the trophic state ranging from oligotrophic to hyper-

trophic. Overall, EC50ranged from 260 to 500 lM for diatoms in so-called oligo- and mesotrophic regimes, from 560 to 1500 lMfor cyanobacteria in eutrophic regimes, and from 1500 to

18,500 lM for chlorophytes in hypertrophic regimes. Quantumyield assesses short-term effects, but nevertheless, it reflects the

longer term effects depicted inTable 2.

Other variables, such as the Photosynthetically Available Radia-

tion (PAR) could modulate the toxic effects of high ammonium lev-

els 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,Admi-raal (1977)reported that inhibition of photosynthesis of estuarine

benthic diatoms growing on 500 lM ammonium was enhanced athigh PAR. This was confirmed by Hillebrand and Sommer (1996)

who found 300 lM to be toxic to Pseudo-nitzschia multiseries at230lmol photons m2 s1, but not at 25 lmol photons m2 s1.On shorter time scales (minutes to hours),Drath et al. (2008) re-

ported more inhibition of PS II activity in Synechocystis at 40 than

at 10 lmol photons m2 s1. The sensitivity ofMicrocystis aerugin-osato ammonium increased with PAR in the range 501000 lmolphotons 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 cul-

ture collection.Berge et al. (2012) tested species that were main-

tained 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 simulatedenvironmental factors to ensure applicability to field observations.

D

inophyceae

A.

minutum

100

20

NA

NA

6200

NA

5

96

NA

Changand

McClean(1997)

A.

minutum

50

20

NA

NA

6200

NA

5

77

NA

Changand

McClean(1997)

A.

minutum

25

20

NA

NA

6200

NA

5

30

NA

Changand

McClean(1997)

Gym.

splendens

200

34020

NA

NA

25200

NA

8

753(lag

)NA

NO3

Thomaset

al.(1980)

Gyr.instriatum

150

20

NA

NA

5250

NA

8

NS

NA

Nagasoeet

al.(2010)

K.

mikimotoi

100

21

8.0

NA

5500

0.262

8

234

9

NO3

Suksomjitetal.(2009a)

L.

polyedrum

200

20

NA

NA

50200

NA

5

126

NA

NO3

Thomaset

al.(1980)

Peridinium

sp.

100

18

6.0

7.2

1010,0

00

03

4

2264

1

Barker(193

5)

P.

micans

100

18

6.0

7.2

1010,0

00

03

4

2674

1

Barker(193

5)

P.

micans

NA

NA

8.1

7.7

74000

NA

6

753

NA

Zgurovskay

aandKustenko(1968)

R

aphidophyceae

Chattonellaantiqua

100

21

8.0

NA

5500

0.262

7

489

19

NO3


H.

carterae

160

15

NA

NA

27875

NA

6

379

NA

Changand

Page(1995)

H.

carterae

80

15

NA

NA

27875

NA

6

183

NA

Changand

Page(1995)

H.

carterae

40

15

NA

NA

27875

NA

6

101

NA

Changand

Page(1995)

H.

akashiwo

100

21

8.0

NA

101500

0.462

10

556

28

NO3




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Przytocka-Jusiak et al. (1977)acclimated an ammonium sensi-

tive strain of Chlorella vulgaris originally maintained at 9860 lM,by culturing in medium containing inhibitory (53,570lM), butnot 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 exces-

sively 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 (84) 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 respectivelyand 6.37.9 at higher ammonium concentrations between 54,570

and 142,860 lM).

The reverse acclimation (called de-adaptation by Przytocka-

Jusiak et al. (1977)) was also carried out in the same study using

6 successive transfers in a medium with 9860lM of either ammo-nium, nitrate or urea. The acquired tolerance to high ammonium

was found to be stable for at least 30 days, with a slight (624h)

time lag for cells pre-conditioned on nitrate or urea. Since 30 days

could represent about 50 cell generations, adaptation rather than

acclimation might have been involved.

2.2. Influence of high ammonium levels on phytoplankton

communities

Ammonium concentrations up to 150lM in coastal or estuarinewaters(Admiraal, 1977;Thomas et al., 1980;Han et al., 1992;Suksomjit

etal., 2009a,b;Yamamoto etal., 2010) may betoxicto manydiatomspe-

cies such asChaetoceros (Suksomjit et al.,2009b),Coscinodiscus (Guillard,

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 reachVmaxfor

ammonium uptake. PAR inlmol photons m2 s1. NA = not available. Nat: natural. Lim: limited; starv: starved; suff: sufficient.

Species PAR T(C) pH Previous N source Physiological

status

NH4addition (lmol N/l) Lag (h) References

Chlamydomonas reinhardii 65 25 7.0 NO3 N starv. 16 h 900 0 Thacker and Syrett (1972)

Chlamydomonas reinhardii 65 25 7.0 NH4 N suff. 200 0 Cullimore and Sims (1980)

Chlamydomonas reinhardii 500 25 7.0 NH4 N suff. 200 0 Florencio and Vega (1983)

Chlorella fusca 70 25 7.4 NH4NO3 N starv.16 h 9000 48 Syrett and Morris (1963)Chlorella fusca 140 25 7.4 NH4NO3 N starv. 16 h 9000 24

Chlorella vulgaris Dark 25 6.0 NH4NO3 N starv. 16 h 10,000 0 Syrett and Fowden (1952)

Ditylum brightwellii 175 15 NA NO3 N.A. 1 75 Eppley et al. (1969)

Emiliana huxleyi 150 13 NA NO3 N suff. 5 85 Page et al. (1999)

Emiliana huxleyi 150 13 NA NO3 N suff. 10 60

Emiliana huxleyi 150 13 NA NO3 N suff. 17 72

Lingulodinium polyedrum 175 15 NA NO3 NA 3 24 Eppley et al. (1969)

Lingulodinium polyedrum 100 18 NA NO3 N starv. 24 h 10 0 Harrison (1976)

Haslea ostrearia 100 NA 7.8 NO3 N suff. 30 48 Robert and Maestrini (1986)

Nitzschia ovalis 100 NA 7.8 NO3 N suff. 30 48 Robert and Maestrini (1986)

Nitzschia ovalis nat PAR NA NO3 N starv. 24 h 40 24 Maestrini et al. (1986)

Phaeodactylum

tricornutum

NA NA 7.6

8.0

NO3 N suff. 200 96 ZoBell (1935)

Phaeodactylum

tricornutum

200 20 8.0 NH4 N suff. 300 0 Cresswell and Syrett (1982)

Platymonas striata 90 20 NA NH4 N starv. 1 h 1200 1 Ricketts (1988)

Skeletonema costatum 310 17 NA NA N suff. 5 0 Conway (1977)Skeletonema costatum 490 17 NA NH4 N lim. 8 0 Conway (1977)

Skeletonema costatum NA 13 NA NO3 N starv.


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1963),Cyclotella(Guillard, 1963; Tadros and Johansen 1968), Nitzschia

seriata (Guillard, 1963), Pseudo-nitzschia multiseries (Hillebrand and

Sommer, 1996), some species ofSkeletonema (Guillard, 1963), coastal

clones ofThalassiosira pseudonana (Guillard,1963). SkoglundandJensen

(1976) found that S. costatum, Thalassiosira pseudonana and Phaeodacty-

lumtoleratedammoniumconcentrationsto>450lM.Incontrast,higherammoniummayactually stimulategrowthof other phytoplankterssuch

asSynechococcus

(Birdsey and Lynch, 1962; Kapp et al., 1975; Neilsonand Larsson, 1980; Dai et al., 2008).

Suksomjit et al. (2009b)compared the ammonium tolerance of

Skeletonema isolated from the Seto Inland Sea (low ambient ammo-

nium) with an isolate from Dokai Bay (ammonium 100lM) andfound that growth rate decreased at 500 lM in the former, but theDokai Bay isolate tolerated ammonium up to 1500 lM. Thissuggests thatSkeletonemafrom Dokai has adapted to the continu-

ally high ammonium concentrations of100lM in Dokai Bay.This is somewhat similar to the study ofHan et al. (1992)who

found that the volume specific photosynthetic rate ofS. costatumin

Tokyo Bay reached maximal values at around 140 lM ammoniumand decreased at higher concentrations. Takahashi and Fukazawa

(1982) using natural communities from Osaka Bay (Japan), re-

ported that 50 lM enrichment stimulated the growth ofEutreptiel-la 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 10lM, while thereverse was true for Gymnodinium sp. (0.15 vs. 0.45 d1) and

Heterosigmasp. (0.5 vs. 0.3 d1).

Domingues etal.(2011) reportedspecificnetgrowth ratesfor several

classes of phytoplankton in a tidal estuary following differential enrich-

ments with ammonium (from 1 to 100 lM) and nitrate (100lM) overseveralseasons.Overall,green algae growth wasconsistently stimulated

by ammonium additions (up to 100 fold)). Cyanobacteria (reported in

summeronly) growth rateswere stronglyenhancedby highammonium

(up to 8-fold increase). Growth rates of diatoms decreased significantly

inspring andsummerwithammoniumadditions>50lM,butthis couldbe due to silicate limitation (estimated from data inDomingues et al.,

2010). Dinoflagellates were clearly inhibited by 100lM ammonium in

the springsummer transition and in summer (even reaching negativegrowth 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,Don-

ald 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 toxicto phytoplankton assemblages with reductions in algal biomass

and species richness (Livingston et al., 2002). Microcosm experi-

ments with the dominant diatom Skeletonema showed inhibitionbetween 7 and 20lM (i.e. surprisingly low compared to otherstudies) 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 dis-

tribution of benthic diatoms on an estuarine mudflat (Admiraal

and Peletier, 1980). Navicula salinarum and Gyrosigma fasciola tol-

erated ammonium concentrations of 700010,000 lM, while fourother species were less tolerant at 20007000 lM.

3. Acclimation time for ammonium uptake at high ammonium

concentrations

In the previous section, we reviewed the effect of ammoniumon 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 trans-

late 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 acclima-

tion period for ammonium uptake at high ammonium concentra-

tions (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 7585 h in

the early study by Syrett and Morris (1963) on Chlorella fusca(Mor-

ris, 1974). Possibly, the large amount of unbuffered ammonium

chloride added (9000 lM) led to an immediate decrease in theambient pH (Collos et al., 1992) that was deleterious to the initial

uptake of ammonium or to growth in general. This time lag prob-

ably 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 (18lM) still led to 24 htime 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 precon-

ditioning 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 spe-

cies 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 shutdown surge uptake.

3.1. Transition phases of uptake rates at high ammonium

concentrations

Table 4summarizes the data on phase transitions (i.e. imple-

mentation of a second higher rate uptake system at higher

substrate concentrations). This phenomenon of bi-phasic or multi-

phasic uptake systems is well known in higher plants (Clarkson

and Lttge, 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 concentra-

tions) have been calculated from available published data. Such

ratios in uptake are variable, but in some cases they can reach veryhigh 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, dinoflag-

ellates) 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 ofAlexandrium

catenella (Collos et al., 2004), Chlorella vulgaris (Urhan, 1932), T.

weissflogii (Conover, 1975), Microcystis (Dai et al., 2008) and Nanno-

chloropsis (Hii et al., 2011). This enhanced uptake at high ammo-nium concentrations enables some species to acquire the limiting

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nutrient faster than other species and enables them to exploit tran-

sient 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

Lttge, 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 cells metabo-

lism. 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 forA. catenella (Collos et al., 2006) and up to66% ofthe cel-

lular N content for T. weissflogii (Conover, 1975) without any

apparent detrimental effect on cell growth. At typical cytoplasmic

and vacuolar pHs of around 78 (Altenburger et al., 1991and ref-

erences therein; Taylor et al., 2012), ammonium would be the

dominant form rather than ammonia.

From an energetic point of view, ammonium should enhancegrowth relative to nitrate, especially at low light. However, con-

trary to expectations, Thompson et al. (1989)showed that higher

growth rate only occurred at high light where energy is not limit-

ing. As suggested by Collier et al. (2012), it is possible that the pre-

dicted lower metabolic cost of ammonium compared to nitrate in

terms of only N assimilation is offset by higher costs in terms of re-

pair 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 re-

sult 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 ex-

plored 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 respira-

tion upon addition of ammonium to nitrogen starved cells, andthe 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 CO2produced dur-

ing 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 evi-

dence 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 suppres-

sion 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 isthen 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 impor-

tance in the interpretation of such changes in carbon fixation re-

sults 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 byParker et al. (2012b)

led to the development ofS. costatum with an initial ammonium le-

vel of about 10 lM, and this is consistent with the relatively highEC50 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 re-

mains unexplained so far. Thus, it is unlikely that the 24 h incuba-

tions 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

chlorophylla (chla) content ofProrocentrum micansdid not change

between 1 and 71lM NH4, that ofS. costatum decreased by a factorof 2.5 between 1 and 71 lM and by a factor of 9 between 1 and714lM (Zgurovskaya and Kustenko, 1968), in parallel with de-creases in photosynthesis (oxygen evolution). This was similar to

the response ofNostocsp. where a 10-fold decrease in chlacontent

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 to3 pg/cell as ammonium further increased from 1000 to

32,000lM, without effect on growth (Fabregas et al., 1989). Thisindicates that there are compensatory mechanisms to counteract

detrimental effects on pigments. InChlorella vulgaris, the chlacon-

tent increased continuously from 0.1 to 2.1 pg/cell) between 714

and 71,428lM (Tam and Wong, 1996), and also in Chlamydomonasreinhardtii (from 1.1 to 3.9 pg/cell) between 100 and 10,000lM

(Giordano et al., 2003). Thus, there are clear detrimental effectsof high ammonium levels on the chlacontent of diatoms and cya-

nophyceae, that are also reflected in reduced growthrates (Table 2),

but none in Chlorophyceae. In contrast, Leong and Taguchi (2004)

reported an increase (4-fold) in chla /cell between 6 and 100 lMammonium, 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 up-

take (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 med-

ia 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 20004000lM and large scaleforest decline has been linked to anthropogenic ammonium inputs

and soil acidification (Pearson and Stewart, 1993; Britto and Kron-zucker, 2002). Reasons for ammonium toxicity include, proton

16 Y. Collos, P.J. Harrison/ Marine Pollution Bulletin 80 (2014) 823


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extrusion, cytosolic pH disturbances displacement of crucial cat-

ions, and shifts in carbohydrate status. Sensitive plants such as bar-

ley show chlorosis of the leaves, significantly reduced growth, a

decrease in cellular cations and a marked acidification in the

hydroponic medium at 10,000lM vs. 100 lM. Britto et al.(2001a,b) showed that at high ammonium concentrations, barley

roots experience a breakdown in the regulation of ammonium in-

flux, 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 poten-

tial 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 NH4that 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 con-

centrations. In higher plants, it is not known why ammonium tox-

icity 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 mil-

limolar range (Britto et al., 2001a). Thus, ammonium toxicity may

be alleviated by converting it quickly to amino acids (e.g. gluta-

mine/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 com-

pounds to avoid toxicity (Giordano et al., 2003). In addition to

the reduction in cytosolic ammonium by various synthesis path-

ways, ammonium may be transported to the vacuole where acid

trapping of ammonium may occur. Ammonium may also be re-

moved 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 massmortality 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 anthropogenicinputs and algal biomass decomposition. Brun et al. (2002) as-

sessed the toxicity of ammonium pulses on the survival and

growth ofZostera noltii and found that the toxic effect depended

on the internal carbon balance between photosynthesis and su-

crose 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 car-

bon demands arising from ammonium assimilation. Brun et al.(2002)found that a repeated large pulse of ammonium quickly in-

creased nitrogen transport and assimilation, and carbon demands

and drained the carbon pool, resulting in intracellular accumula-

tion of ammonium and hence toxic effects to photosynthesis and

growth. At pH 8,Zostera marina showed no effects of high ammo-

nium concentrations up to 150lM, but at pH 9, it became necroticin a few days (van der Heide et al., 2008). Remarkably, with high

shoot densities at pH 9, there was no necrosis and toxic effects ap-

peared to be alleviated through joint ammonium uptake occur-ring 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 125lMand even down to 25 lM. They found that low light enhancedammonium toxicity because photosynthesis was reduced along

with carbon stores that are needed for the assimilation of ammo-

nium into amino acids. Similarly, Villazn 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 tem-

peratures (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 sur-

rounding seawater pH is8.2, compared to pH of 7.5 in sediments,

despite the much higher ammonium concentrations in sediments

of between 1001000 lM.In a comparative study of two duckweeds,Monselise and Kost

(1993) found that Lemna gibbafrom a high ammonia site was more

efficient at detoxifying excessive ammonia by simply accelerating

the usual GS-GOGAT nitrogen assimilation system (i.e. an ammo-

nium trapping process), producing glutamine and c-aminobuty-rate and leaving little intracellular free ammonia/ammonium. In

contrast,Wolffia arrhizafrom a low ammonium site had a less effi-

cient ammonium assimilation system, probably due to a relative

shortage of glutamate. This suggestion was supported by the addi-

tion ofa-ketoglutarate which resulted in a significant decrease inintracellular ammonia. Thus, a-ketoglutarate is a precursor for

the ammonium trapping mechanism. Substantial evidence indi-cates 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, crusta-ceans 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 Melbournes sewage treat-

ment plant caused a decrease in brown macroalgae and an increase

in opportunistic green macroalgae near the discharge site. Subse-

quent 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 efflu-

ent 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). Com-parisons of toxicity using calculated unionized ammonia (NH3) and

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total ammonia gave similar results. High ammonium toxicity val-

ues for four marine amphipods ranged from a EC50 of 3000 to

10,000lM, probably because amphipods are exposed to higherammonium concentrations associated with the sediments as com-

pared to the water column (Kohn et al., 1994).

In a very carefully designed bioassay, Kllqvist and Svenson

(2003) used the chlorophyte Nephroselmis pyriformis since it was

the most sensitive of the nine species that they tested. They used3 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 in-

ferred that the toxicity was mainly due to ammonia (e.g. Hogan

et al., 2005; Adams et al., 2008). Kllqvist and Svenson (2003)

found that ammonia had an EC50 of 2.4 lM, while the EC50 forammonium was 100 times higher at 224 lM and the EC50for to-tal ammonia at pH = 8.0 was 71lM. They concluded that there is ajoint 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 Aus-

tralian and New Zealand guidelines (termed trigger values) for

marine and estuarine waters are as follows : 65 lM total ammoniafor 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 USEPAhas 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 ex-

panded toxicity data set and their new 95% species protection va-

lue was 32.9 lM, about half the previous value of 65lM. For

sediment pore waters, a trigger value of 300lM was recom-mended. They stated that a guideline concentration of 565lM(i.e. the concentration below the EC50for 95% of the species), rep-

resents 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 whenthe 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 ammo-

nium concentrations for the more sensitive fish. Nannochloropsis

sp. appears to be an ideal species since it can tolerate up to

900lM ammonium and grows significantly faster on ammoniumthan nitrate (Hii et al., 2011). In fish ponds, there is a daily pHcycle

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 in-

crease in the ammonia concentration. Ammonia may be produced

in/near the sediments due to the decomposition of the algae andthe 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 NO2 and NO

3 . Ammonia concentra-

tions tend to be higher in the winter (180 to 285 lM) in the fishponds and lowest in summer (35lM) due to the highest algalgrowth in summer (Hargreaves and Tucker, 2004). The complex

nitrogen biogeochemistry of aquaculture ponds has been well re-viewed byHargreaves (1998).

A common problem in many ponds is the occurrence of the tox-

in-producerPrymnesium parvum. For some fish, ammonium sul-

phate can be added to an ammonia concentration of 1020 lMthat will killP. 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 comparedto 130 lM for marine fish, indicating that marine species areslightly 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 waste-

water and concentrations range from 10200 mg L1 (1000

20,000lM) (Konig et al., 1987; Thomas et al., 1980). There hasbeen considerable interest in using high rate oxidation ponds to

treat wastewater (Abeliovich and Azov, 1976). One alga that is par-

ticularly 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

20,000lM (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 (58 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, 20003000 lM total ammonia led to a 5090% reduc-tion 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 ammo-

nia 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

forPhaeodactylum tricornutumwhich grew at a pH of 10.3. The un-ique ability ofP. tricornutumto grow at high pHs is a major factor

that explains its frequent dominance in large-scale outdoor cul-

tures (Goldman et al., 1982c). Therefore,P. tricornutum would ap-

pear to be an ideal/unique species for further studies on the

interaction of pH and ammonium toxicity. Of the freshwater spe-

cies, Scenedesmus obliquus grewbetter at a pHof 10.6 than Chlorella

vulgaris.

Duckweed systems are another option for sustainable wastewa-

ter treatment because they have high growth rates and high nutri-

ent content that can be used for animal feeds ( Krner et al., 2001).

At low total ammonia concentrations of


12/16

(Spirodela polyrrhiza),Caicedo et al. (2000) found that both N forms

caused growth inhibition, but ammonia inhibition occurs at much

lower concentrations than ammonium inhibition. An increase in

pH can produce other cellular effects besides increasing the ammo-

nia/ammonium ratio and the use of buffers was recommended in

order to separate the pH effects from the total ammonia concentra-

tion effect (Caicedo et al., 2000).

6. Conclusions

The effects of ammonium on microalgae can occur on two dif-

ferent 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 since a tran-

sient lag/induction/acclimation period may not influence the out-

come of species competition in the field that is expected to occur

over several days of growth.

In spite of large intra- and inter-specific variability, there aresignificant differences between classes of unicellular algae with re-

spect to the effect of high ammonium levels on growth rates. Lab-

oratory experiments indicate that growth rates of chlorophytes

followed by diatoms were the most tolerant and dinoflagellates

were the most sensitive to high ammonium. With EC50 values

ranging from 30 to 2700 lM for dinoflagellates, 604000 lM fordiatoms and up to 380056,000 lM for chlorophytes, and mostfield ammonium concentrations well below 100 lM, ammoniumtoxicity effects on growth rates are not likely to occur in the field.

In addition, toxicity occurs more frequently at high pHs >9 where

ammonia is more abundant, and these high pHs are seldom ob-

served in the field. The ranking of algal groups observed in the cul-

tures is somewhat reproduced in the field, although such

comparisons are fraught with difficulties such as long-term accli-mation 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 byKllqvist

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 impor-

tant 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.

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Cita [6] Aclimatación y toxicidad de las altas concentraciones de amonio.

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